NUREG/CR-6910PNNL-15872

Alternative Conceptual Modelsfor Assessing Food-ChainPathways in Biosphere Models

Pacific Northwest National Laboratory

U.S. Nuclear Regulatory CommissionOffice of Nuclear Regulatory Research .Washington, DC 20555-0001 A&

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NUREG/CR-6910PNNL-15872

Alternative Conceptual Modelsfor Assessing Food-ChainPathways in Biosphere Models

Manuscript Completed: June 2006Date Published: June 2006

Prepared byB. A. Napier

Pacific Northwest National LaboratoryRichland, WA 99352

P. R. Reed, NRC Project Manager

Prepared forDivision of Fuel, Engineering, and Radiological Research

Office of Nuclear Regulatory ResearchU.S. Nuclear Regulatory CommissionWashington, DC 20555-0001Job Code Y6469

Abstract

This report describes alternative approaches to modeling the key processes inradionuclide transport in the biosphere. Because the focus of the NRC projectAssessment of Food Chain Pathway Parameters in Biosphere Models is the food-chainmodels used in performance assessments of radioactive waste disposal facilities, modelsand approaches applicable over relatively long periods (more than one year) areevaluated, as opposed to approaches detailing radionuclide behavior over the shorterperiods applicable to acute, accident-type, calculations. There are a number of importantfeatures and processes that all terrestrial biosphere models must address: these includeradionuclide behavior in soils, interception of deposition onto vegetation, weathering ofintercepted material from plant surfaces, foliar absorption and translocation within plantsto other vegetative structures, uptake from soil by plant roots, and transfer from plants toanimals and animal products.

The report provides a detailed discussion of possible alternative approaches tomodeling the food-chain pathway in biosphere models and recommends the following:

" For soil, an annual average model incorporating:o Accumulation from irrigation or atmospheric deposition;o Leaching to deeper soil, using a Kd-modulated leach rate;o Harvest removal, averaged over crop typeso Uniform mixing in a reasonable (15-25 cm) upper soil surface layer,

implicitly caused by plowing, bioturbation, and leaching;o Neglecting radionuclide fixation, because appropriate long-term

measurements of distribution coefficient and concentration ratio willalready incorporate the effects;

o Neglecting surface-soil erosion losses, because eroded material from onelocation may accumulate in another, cancelling any perceived benefit.

" For resuspension, a mass-loading approach, because it has the lowest variabilityand is the most easily defended.

" Forfoliar interception, models incorporating:o Dry interception following the basic model suggested by Chamberlain

(1967) as updated by Adriano et al. 1982;o Wet interception considering one of the algorithms derived from the data

of Hoffman et al. (1989) or Prohl and Hoffrman (1993)." For plant contamination with soil, an adhesion model applied so that material

translocated is not subject to weathering (e.g., IAEA 2003; Wu 2003)." For weathering, a single exponential model with a half-time between about 10

and 20 days, applied to material on the surface of the plant only." For translocation, a radionuclide-specific (or chemical-class-specific)

implementation allowing translocated materials to avoid being removed viaweathering processes (e.g., IAEA 2003; Wu 2003).

" For soil-to-plant transfer, a radionuclide- and plant-type-specific concentrationratio, where available, with:

o Site-specific application of additional crops, such as mushrooms;

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o The specific-activity model of Peterson and Davis (2002) for elementaland oxide forms of tritium;

o A general specific-activity model for 14C in air, adapted to the RESRADfamily of codes (Yu et al. 2001) model for transfer from irrigation water toair as 14C0 2;

o Perhaps site-specific application of specific-activity models for iodine andother micro-nutrient and macro-nutrient elements and their chemicalanalogues. This is a simplification of the need to allow for homeostaticregulation of some elements.

* Until a simple, robust Fruit Tree model is developed, continue to use theconcentration ratio approach for fruits.

" For animalproducts, a transfer factor approach applied to average daily feed,water, and soil intake by the animal where data are available.

o For element/animal combinations where data are lacking, a justified mixof specific-activity and allometric models should be considered.

" Forfoodprocessing, because the data are sparse and the reduction is generallysmall, modelers are justified in ignoring losses (which is equivalent to using afood-processing transmission factor of 1.0).

This type of information is directly useful in formulating inputs to radioecological andfood-chain models used in performance assessments and other kinds of environmentalassessment. This food-chain pathway data may be used by the NRC staff to assess doseto persons in the reference biosphere (e.g., persons who live and work in an areapotentially affected by radionuclide releases) of waste disposal facilities anddecommissioning sites.

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FOREWORD

The food chain pathway contributes to the potential dose received by members of the public asa result of the potential release of radionuclides to the environment from many facilities licensedby the U.S. Nuclear Regulatory Commission (NRC). To quantify the contribution of thispathway, NRC developed performance assessment strategies involving biosphere computercodes to evaluate the potential dose to humans. These biosphere codes incorporate parametersfor radionuclide uptake in plant roots and leaves, as well as animal products, to aid in predictingthe radionuclide concentrations that humans would ingest in the event of an environmentalrelease.

This report describes alternatives to current approaches for modeling key features andprocesses to assess radionuclide transport and behavior in food-chain pathways in thebiosphere. Section 2 describes some important concepts, features and processes that allbiosphere models must address. Section 3 provides descriptions of various ways of modelingthese important features and processes. Finally, Section 4 discusses the types of modelingapproaches that are most applicable to assess doses to persons who live and work in areasaffected by potential radionuclide releases from nuclear facilities.

Included in this report are detailed discussions on approaches to modeling long-lived radionuclidebehavior in soils in and near the plant root system, soil resuspension, and wet and dry leafinterception of deposition onto vegetation. There are also discussions on weathering ofintercepted material from plant surfaces, foliar absorption and translocation within plants to othervegetative structures, uptake in fruit and nut trees, radionuclide uptake from soil by plant roots,and the transfer of radionuclides from plants to animals and animal products.

For some of the processes described in this report, several methods of different levels ofcomplexity are described depending on the type of analyses being performed. For otherprocesses, limited information was available and the report includes only one or two altemativeapproaches to the calculations.

NRC is currently using the results described in this report to interpret the results of ongoingexperiments at Pacific Northwest National Laboratory to determine radionuclide uptake factorsin plants and animals. For the future, the Commission expects to use the data and informationpresented in this report to reduce uncertainties in food-chain-pathway dose assessments. Theresults of the biosphere research prograrh improve the NRC staff's understanding of the featuresand process that affect estimates of dose from important long-lived radionuclides in the food-chainpathway.

Brian W. Sheron, DirectorOffice of Nuclear Regulatory Research

v

Contents

Abstract .............................................................................................................................. iii

Foreword ............................................................................................................................. v

List of Figures .................................................................................................................. viii

List of Tables ..................................................................................................................... ix

Acknowledgements ............................................................................................................ x

1.0 Introduction .............................................................................................................. 1-1

2.0 Concepts in Biosphere Modeling .................................. 2-1

3.0 M odeling Key Processes .......................................................................................... 3-13.1 M odeling Radionuclides in Soil ............................................................................ 3-13.2 M odeling Resuspension ........................................................................................ 3-83.3 M odeling Radionuclides in Vegetation ............................................................... 3-10

3.3.1 M odeling Foliar Interception and Retention .......................................... 3-113.3.2 M odeling Contamination of Plant Surfaces with Soil ........................... 3-163.3.3 M odeling W eathering from Plant Surfaces ............................................ 3-173.3.4 M odeling Foliar Absorption and Translocation ..................................... 3-193.3.5 Modeling Soil-to-Plant Uptake ....................... .... 3-22

3.4 M odeling the Uptake of Radionuclides in Animal Products .............................. 3-413.4.1 Feed to Animal Product Transfer Factors .............................................. 3-4 13.4.2 The Observed Ratio in Animal Products ............................................... 3-423.4.3 Unusual Animal Products ........................... ; .......................................... 3-433.4.4 Specific Activity M odels for Animal Products ...................................... 3-443.4.5 Kinetic M odeling for Animals ............................................................... 3-463.4.6 Allometric M odels for Uptake by Animals ........................................... 3-50

3.5 Modeling the Effects of Food Processing on Radionuclide Concentrations ....... 3-513.6 M odeling Other Terrestrial Exposure Pathways ................................................. 3-56

3.6.1 Unique Direct Pathways ........................................................................ 3-563.6.2 Unique Indirect Pathways ...................................................................... 3-57

4.0 Recommendations .................................................................................................... 4-1

5.0 References ................................................................................................................ 5-1

vii

List of Figures

Figure 1. Idealized Behavior of Radionuclide Concentration in Soil FollowingLong Periods of Irrigation for Mobile and Immobile Radionuclides ............ 3-7

Figure 2. A Basic Model for Contamination of Plants ................................................ 3-11Figure 3. Relationships are shown between tissue concentrations and

environmental concentrations for non-nutrient and nutrient compounds... 3-33

viii

List of Tables

Table 1. Biotic Transport Quantities (adapted from McKenzie et al. 1986) ............... 3-8Table 2. Probable Bioavailability/Cuticular Transfer of Selected Elements ............. 3-21

Table 3. Aggregate transfer coefficient for cesium for 4 mushroom classes ............. 3-32Table 4. In a biological classification of metals, shaded cells are non-nutrient

analogs; bold cells are nonessential metals .................................................. 3-34

Table 5. Weight Fractions of Hydrogen and Carbon in Environmental Media,Vegetation, and Animal Products ................................................................. 3-36

Table 6. Parameter values for calculating HTO and OBT concentrations in plantproducts ........................................................................................................ 3-37

Table 7. Suggested parameter values for calculating HTO and OBTconcentrations in animal products ............................................................... 3-46

Table 8. Recommended Values of Food Processing Retention Factors for MostContam inants ............................................................................................... 3-55

Table 9. Recommended Values of Food Processing Retention Factors for Tritium. 3-55

ix

Acknowledgements

The author is particularly grateful for the technical guidance, review, and encouragementprovided by P. R. Reed of the U.S. Nuclear Regulatory Commission. The author thanksW.E. Kennedy, Jr. (Dade Moeller and Associates), K.A. Higley (Oregon StateUniversity), F.O. Hoffman and J. R. Trabalka (SENES Oak Ridge), C.A. Brandt(PNNL), and NRC staff for detailed review comments on the draft of this report.

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1.0 Introduction

The U.S. Nuclear Regulatory Commission's project Assessment of Food Chain PathwayParameters in Biosphere Models has been established to assess and evaluate a number ofkey parameters used in the food-chain models used in performance assessments ofradioactive waste disposal facilities. The objectives of the research program include:

" Provide data and information for the important features, events, and processes ofthe pathway models for use in biosphere computer codes. These codes calculatethe total effective dose equivalent (TEDE) to the average member of the criticalgroup and/or reasonably maximally exposed individual, for example, fromradionuclides in the contaminated ground water release scenarios in NRC'sperformance assessments of waste disposal facilities and decommissioning sites,

" Reduce uncertainties in food-chain pathway analysis from the agriculture scenariosof biosphere models in performance assessment calculations,

" Provide better data and information for food-chain pathway analyses by:o Performing laboratory and field experiments, including integral and

separate effect experiments, to evaluate the potential pathways and uptakemechanisms of plants and animals contaminated by long-livedradionuclides,

o Presenting food-chain pathway data and information by regional and localgeographical locations,

o Quantifying uncertainties in the radioactive contamination of food cropsand long-term build up of radionuclides in soils with contaminated groundwater from water irrigation systems,

o Determining data on factors affecting radionuclide uptake of food cropsincluding irrigation water processes, soil physical and chemical properties,soil leaching and retention properties near crop roots, soil resuspensionfactors and other soil and plant characteristics.

The results of this research program will provide needed food-chain and animal productpathway data and information for important radionuclides that may be used by the NRCstaff to assess dose to persons who live and work in areas potentially affected byradionuclide releases from waste disposal facilities and decommissioning sites.

The biosphere model is the last model in a series of models used by the NRC staff inperformance assessments of waste-disposal facilities and decommissioning sites. Thebiosphere models used by the NRC staff in its performance assessments are based onconceptual models, their mathematical representation, and the implementing computercodes. The biosphere models used by the NRC, and their implementing computer codes,include hypotheses, assumptions, and simplifications that describe the referencebiosphere in the vicinity of the waste-disposal facilities and decommissioning sites.These computer codes consider (1) radionuclide transport through many food-chainpathways such as irrigation-water deposition on vegetation and soil surfaces, (2) cropinterception and retention, (3) radionuclide buildup in soils as a result of long-term

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irrigation deposition, (4) soil radionuclide leaching and retention mechanisms in plantroot zones, (5) resuspension of radionuclide-contaminated soil onto vegetation, andtransfer via gaseous phase radionuclides arising from soil, (6) soil to plant uptake viaroots, (7) uptake by stems/leaves of soil-derived gaseous contaminants, (8) feed-to-animal product transfer, and (9) individual food product consumption rates. Biospheremodels used in performance assessments provide the long-term, time-dependentconcentrations of radionuclides in soil, plants, and animal products that could beconsumed by future human residents of the biosphere. When coupled with projections ofhuman behavior, they also provide estimates of annual or lifetime radiation dose; thatcomponent of the biosphere modeling is not addressed in this report.

Section lof this report is this introduction.

Section 2 of this report describes some key concepts in biosphere modeling - important.features and processes that all biosphere models must address in some manner. Forterrestrial models, these include radionuclide behavior in soils, interception of depositiononto vegetation, weathering of intercepted material from plant surfaces, foliar absorptionand translocation within plants to other vegetative structures, uptake from soil by plantroots, and transfer from plants to animals and animal products.

Section 3 of this report provides descriptions of various ways of modeling theseimportant features and processes. For some processes, various methods of differentlevels of complexity are applied depending on the type of analyses being performed. Forother processes, very limited information allows only one or two alternative approachesto the calculations.

Section 4 makes recommendations on the types of modeling approaches that are mostapplicable for models to be used to assess doses to persons who live and work in areaspotentially affected by radionuclide releases from waste disposal facilities anddecommissioning sites.

The results are expected to be useful in:

" supporting the development of regulatory criteria (e.g., guidance, technicalpositions) for food-chain pathway issues involving biosphere models

" providing a basis for evaluating and auditing an applicant's or licensee'sbiosphere and food-chain pathway data, information, analyses, conceptualmodels, and computer codes used in license submittals

" providing NRC staff with data and information for resolving biosphere issuesinvolving irrigation pathways, food and animal transfer factors, and groundwaterradionuclide release scenarios.

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The results of the research program improve the NRC staffs understanding of thefeatures and processes for some important long-lived radionuclides in biospheremodeling of the performance-assessment process.

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2.0 Concepts in Biosphere Modeling

There are a number of important features and processes that all biosphere models mustaddress in some manner (UNSCEAR 2000). For terrestrial models, these includeradionuclide behavior in soils, interception of deposition onto vegetation, weathering ofintercepted material from plant surfaces, foliar absorption and translocation within plantsto other vegetative structures, uptake from soil by plant roots, and transfer from plants toanimals and animal products. Because different models and model types approach thesefeatures and processes differently, some of the reasons for the different approaches arediscussed in this section in general terms before specific methods of computation aredescribed.

The primary inputs to estimating radiation doses to individuals and populations are theconcentrations and availabilities of radionuclides in air, water, soil, and foods, and thelevel of exposure of the individuals or groups to each radionuclide in each medium.Because it is not possible to completely characterize the radiological environment, andbecause in many practical cases the concentration of radionuclides in the environmentresulting from reactor operations is too low to measure, mathematical models areemployed to go from what is known to what needs to be known.

Many different types of models have been developed. The various models are directed toanswering different questions. Different models have different levels of detail, differentdegrees of accuracy, and different temporal or spatial scales. Some are used forretrospective analyses (to analyze what has happened - as for determining compliancewith annual radiation dose limits) or for prospective analyses (to project what mayhappen - as for demonstrating compliance with licensing requirements).

Two general classes of environmental transport models have evolved: dynamic (transient)and equilibrium (steady-state). Both tend to describe the environment in terms of various"compartments" such as soil layers, plant types, and animal types (some environmentalmedia may be described in terms of more than one compartment, such as the roots, trunk,branches, and fruit of trees). The dynamic models consider the time-dependent quantitiesof radionuclides in various environmental compartments. The structure of the model isrepresented by a series of coupled differential equations that describe the rate of changeof the amount of the radioactive material in each compartment as a function of thevarious transfer paths into (e.g. irrigation) and out of it (e.g., radioactive decay, harvest).When the equations are evaluated over sufficiently long times with unvarying values ofthe inputs and rate constants, the ratios of the concentrations of the radionuclides in thevarious compartments approach constant values. The system then is considered to be inequilibrium or in a steady state. Dynamic models are frequently used as research tools toinvestigate the detailed phenomena and mechanisms of environmental transport ofradionuclides. For example, they are beneficial in evaluating the consequences of pulseinputs to selected compartments, as might happen following an accidental release.However, because of the temporal resolution demanded of the output, a great deal of

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information is required as input to this type of model, and extensive computer resourcesare required for implementation. By using assumptions of quasi-equilibrium (that is,relatively small changes from year to year in local conditions), the dynamic models maybe simplified into equilibrium models. The equilibrium models lose the ability to answercertain temporally-based questions, but are generally simpler to use, because many of thedetailed rate constants required by the dynamic models can be treated as lumpedparameters. Because most regulations dealing with nuclear waste are written in terms ofannual doses to people, most routine radiological evaluations are performed using theequilibrium models with annual-average values for the input parameters.

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3.0 Modeling Key Processes

Alternative approaches to modeling the key processes in radionuclide transport in thebiosphere are presented in this section. Because the focus of the NRC project Assessmentof Food Chain Pathway Parameters in Biosphere Models is the food-chain models usedin performance assessments of radioactive waste disposal facilities, models andapproaches applicable over relatively long periods (more than one year) are evaluated, asopposed to approaches detailing radionuclide behavior over the shorter periods applicableto acute, accident-type, calculations. International reviews of models of this type thathave been performed in the past decade include the IAEA BIOMASS (AEA 2003)program and the European BIOPROTA program (Albrecht et al. 2005).

3.1 Modeling Radionuclides in SoilSurface soils tend to be compartments for long-term retention and accumulation ofradionuclides released into the accessible environment, either from atmospheric or liquidreleases or from localized depositions. Soils typically are made up primarily of mineralcomponents with additional organic material incorporated from decomposing plantmatter, hosting numerous microorganisms and small invertebrates. The surface layer ofsoils may be mixed by small animals, and in agricultural systems is usually mixedannually by tillage. Radionuclides may become incorporated in vegetation; the mineraland organic constituents of vegetation cycle in the soil and serve as a reservoir as theorganic component decomposes. On small spatial scales, soils may erode, thus movingcontamination from one location to another. Percolating water may transportcontaminants from shallow to deeper soils.

The concentrations, mobility, and bioavailability of radionuclides in surface andsubsurface geologic systems are controlled by numerous hydrologic and geochemicalprocesses. These include hydrologic factors, such as dispersion, advection, and dilution;and geochemical processes, such as aqueous complexation, oxidation/reduction (redox),adsorption/desorption and ion exchange, precipitation/dissolution, diffusion, colloid-facilitated transport, and anion exclusion (Langmuir 1997; Sposito 1989; 1994).Additionally, in the uppermost layer of surface soil, the mobility of radionuclides canalso be increased by biological activity and by the drying and subsequent cracking ofsoils. Colloid-facilitated transport and anion exclusion have received considerableattention recently in that they can enhance the transport of certain radionuclides.However, these processes are hard to quantify, and the extent to which they occur isdifficult to determine. The importance of colloid-facilitated migration, especially inaquifer systems that do not involve fracture flow of groundwater, is still a subject ofdebate.

Some radionuclides, such as technetium, uranium, or plutonium, may be present in morethan one oxidation state in the environment. The adsorption and precipitation behavior ofdifferent oxidation states of a particular radionuclide are usually very different. Forexample, in environmental systems, the most stable oxidation states of technetium are +7and +4 under oxidizing and reducing geochemical conditions, respectively. The chemical

3-1

behavior of technetium in these two oxidation states differs drastically. DissolvedTc(VII) exists as pertechnetate anion, TcO4 , over the complete pH range of naturalwaters under oxic conditions. Because the pertechnetate anion is highly soluble and isnot strongly sorbed, it is highly mobile in most oxidizing systems. Under reducingconditions, however, Tc(IV) exists as hydrolyzed cations and is relatively immobile inthe absence of strongly complexing ligands. Technetium(IV) is highly sorbed and formsthe sparingly soluble TcO2-nH20 solid (Krupka 2003).

Adsorption/desorption (including ion exchange) and precipitation/dissolution areconsidered the most important processes affecting radionuclide interactions with soils.Precipitation/dissolution is more likely to be an important process where elevatedconcentrations of dissolved radionuclides exist, such as in the near-field environment ofradioactive waste disposal facilities or the spill sites of radionuclide-containing wastes orwhere steep pH or redox gradients exist. Adsorption/desorption will likely be the keyprocess controlling radionuclide retardation in areas where trace concentrations ofdissolved radionuclides exist, such as those associated with far-field environments ofdisposal facilities or spill sites or in areas of where soils are to be irrigated usingradionuclide-contaminated water.

The term "sorption" is used as a generic term devoid of mechanism and used to describethe partitioning of dissolved aqueous-phase constituents to a solid phase. When aradionuclide is associated with a geologic material, however, it is usually not known ifthe radionuclide is adsorbed onto the surface of the solid, absorbed into the structure ofthe solid, precipitated as a three-dimensional molecular structure on- the surface of thesolid, or partitioned into the organic matter (Sposito 1989). The term "sorption"encompasses all of the above processes.

The sorption of radionuclides on soils is frequently quantified by the partition (ordistribution) coefficient (Kd). The Kd parameter is a factor related to the partitioning of aradionuclide between the solid and aqueous phases and is defined as the ratio of thequantity of the adsorbate adsorbed per mass of solid to the amount of the adsorbateremaining in solution at equilibrium. Radionuclides that adsorb very strongly to soil havelarge Kd values (typically greater than 100 mL/g) compared to those values forradionuclides that are not significantly retarded by adsorption. Radionuclides that do notadsorb to soil and migrate essentially at the same rate as the waterflow have Kd valuesnear 0 mL/g. The Kd model is the simplest yet least robust sorption model available.However, the Kd metric is the most common measure used in hydrologic transport andbiosphere codes to describe the extent to which contaminants are sorbed to soils. Theprimary advantage of the Kd model is that it is easily inserted into computer codes toquantify the reduction in the extent of transport of a radionuclide relative to groundwater.The Kd is an empirical unit of measurement that attempts to account for various chemicaland physical retardation mechanisms that are influenced by a myriad of variables. Assuch, the KI model is often the subject of criticism (EPA 1999).

Because of recharge from precipitation or irrigation, radionuclides deposited on soils mayleach through the surface layer with percolating water, eventually reaching depths below

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the surface where they are no longer available for uptake by plants and animals. Sorbedmaterials may also travel with soil particles as they are physically transported via wind orwater erosion.

Deposition of radioactive material from the air or irrigation water onto the earth's surfaceis a function of surface area. Atmospheric dispersion model outputs generally deal withdepletion of plumes or puffs of material in terms of mass per unit area per unit time.Irrigation deals with quantities of water per unit area (e.g., acre-feet) per unit time.Although a fraction of depositing activity usually is intercepted by vegetation, it isgenerally assumed that all deposition eventually works its way to6 the soil surface. Theinitial problem is to convert this initially infinitesimally-thin layer into a bulkconcentration in soil, and then to describe its subsequent behavior.

This problem was first addressed in projections of nuclear weapon fallout (Burton 1966;Ng and Thompson 1966). The initial conversion was to assume instantaneous mixing ofdeposition in a soil "root zone" with a depth of 20 cm and a soil density of 2 g/cm 3. Herethe radionuclides remain except for radioactive decay.

This basic concept was adopted in early radiological models (e.g., Soldat and Harr 1971)which used a 15 cm plow depth and resulting area density of 224 kg/m2. This approachwas adopted by the NRC in Regulatory Guide 1.109 (1977), with a soil plow layermixing depth of 15 cm and effective surface density of 240 kg/m2 (equivalent to a soildensity of 1.6 g/cm3). This is assumed to represent the depth to which plowing or otheragricultural practices mix the soil on an annual basis (so that I-year's deposition is mixedinto the rooting zone and available for plant uptake). This became the default approachfor many computer codes written to implement the concepts of this regulatory guide (e.g.,FOOD (Baker, Hoenes, and Soldat 1976), AIRDOS (Moore et al. 1979), RSAC (Wenzel1993)). This basic model became the international standard with the issuance ofInternational Atomic Energy Agency publication Safety Series 57 (IAEA 1982). In theGENII code (Napier et al. 1988), the mixing depth is defined to be a constant 15 cm (6in.). The code MEPAS (Strenge and Chamberlain 1995) allows user inputs, but thedefaults are 15 cm (6 in.) for agricultural soils, and 4 cm (1.6 in.) for residential soils.The PATHWAY (Whicker and Kirchner 1987) model uses 25 cm (10 in.). ECOSYS(Mueller and Prohl 1993) uses 10 cm (4 in.) for pastures and 25 cm (10 in.) for plowedsoil. The NCRP screening model (NCRP 1999) conservatively uses 5 cm (2 in.).Because of its primary use as a code for remediation of previously contaminated soils, theRESRAD family of codes (Yu et al. 2001) default is 200 cm (79 in.). The calculated soilconcentration is inversely related to this assumption. Similarly, the assumed soil densityhas a direct influence on the estimated concentrations.

It is unlikely that long-lived radionuclides introduced into surface soil would remain inone location forever. Therefore, additional processes of radionuclide migration havebeen considered in various models. The first mechanism considered is leaching ofcontaminants out of the root zone and into deeper soils, effectively taking them out of thesystem. This approach is used in models such as GENII (Napier et al. 1988), RESRAD(Yu et al. 2001), the model of Peterson (1983), and PATHWAY (Whicker and Kirchner

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1987). Leaching (as caused by overwatering for example (Schreckhise 1980)) is treatedas an average loss rate term, generally using a first-order rate constant. The leachingterm is thought to be dependent on the type of soil, the radionuclide, and the amount ofwater percolating through the surface and into the deeper soil. The format of Baes andSharp (1981) is frequently used:

P+I-E

= (1)dses, l+&K-i

where X•j = removal rate constant for activity of radionuclide i in the surface soil layer(yr')

P = total annual precipitation (cm/yr)I = total irrigation rate (cm/yr)E = total evapotranspiration rate (cm/yr)ds = depth of soil rooting zone (cm)p, = surface soil bulk density (g/cm3)0, = surface soil volumetric water content (mL/cmi3 )Kd5 i= surface soil distribution coefficient for radionuclide i (mL/g).

In general, because precipitation is variable and because it is difficult to calculate theevapotranspiration, the P+I-E term is approximated as a constant "overwatering" term.

The factor describing mobility (or transportability) of radionuclides in soil is the Kd.This term is empirical and rarely easily available for specific soils in specific locations.Frequently, measurements may be poor; for very insoluble materials, sometimesexperiments may measure precipitation into a solid rather than reversible sorption. Themodels in GENII, in RESRAD, in MEPAS (Strenge and Chamberlain 1995), described inNUREG/CR-5512 (Kennedy and Strenge 1992), and Peterson (1983) all follow the samestructure. (The MILDOS model [Strenge and Bander 1981] is similar, but uses a constant50-year leaching half-time for all contaminants.) No current biosphere models haveaddressed solubility limits, which are considered to be potentially important in somecircumstances (NRC 2002). The IAEA Safety Series 19 model (IAEA 2001) makes adistinction between anionic radionuclides, isotopes of strontium and cesium, and allother radionuclides. Anions such as TcO4 -, cr and F are leached quickly; the defaultvalue for X, is 0.5 yr-I. For strontium and cesium the default value is 0.05 yf-1. For allother nuclides (also non-anionic Tc) the default value is zero.

Some models have slightly more sophisticated surface soil zone models. The ECOSYS,PATHWAY, and MACCS (Jow et al. 1990) models include a second mechanism foreffective removal of radionuclides from the soil for plant uptake. These models includean additional rate constant for "fixation" of contaminants. In these models, theradionuclide is bound to receptor sites in the soil and immobilized. In these models, theradionuclide is no longer mobile, so it is not available for plant uptake, although it stillcontributes to external exposure.

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Measurements of depth profiles of 137 Cs, 239Pu, 241Am, and 37Np in undisturbed soils

from both Chemobyl and weapons' fallout indicate that removal times can be derived thatare considerably faster than those derived from the distribution coefficient Kd. It isinteresting to note that the residence times of the radionuclides investigated are relativelysimilar although their chemical properties are different. Apparently, element independenttransport mechanisms such as the transport of radionuclides attached to clay particles or

bound to soil colloids may also play a role. Since the main transport mechanisms for

strongly bound radionuclides may be to a large extent element-independent, the value for

the residence half-time for the upper soil layer in the European emergency response codeRODOS (Mtller, Gering, and Pr6hl 2004) is applied for all other elements with high Kdas zirconium, niobium, ruthenium, cerium and plutonium - a soil half-life of 100 years is

assumed.

Another potential mechanism for loss of material from the surface soil is harvest removal

(sometimes referred to as cropping), the process by which contaminants leave a localsystem by being harvested with the crop and removed. Hoffmnan and Baes (1979)proposed a linear rate constant to be applied annually as

= B1, MA H. (2)P

where= the harvest removal rate constant (yrf)

Biv = the ratio of concentration of radionuclide i in the harvested plant comparedto that in soil (Bq kg" plant per Bq kg' soil)

Mh = the harvested biomass of vegetation per unit surface area per harvest (kgm"21, the amount of plant material which is removed at harvest and notreturned as recycle

Hn = the number of harvests per year (yfr1)P = the soil area density discussed above (kg in2).

The ERB2A model from the IAEA BIOMASS program (IAEA 2003) uses a similarformulation, including averaging over all crop types with the assumption that long-termagriculture would include crop rotation. A similar expression is used also by Willans(2003) in the MONDRIAN code used at British Nuclear Fuels Limited, as well as by

EPRI (2002). These models assume that, over long periods of time, crops will be rotated,and so each uses the average of the soil and crop parameters for all of the agriculturalpathways assumed.

In the GENII computer codes (Napier et al. 1988; 2002), loss of activity from the surfacesoil zone by harvest is modeled as a step function applied at the end of each calculationalyear. The amount of loss is calculated from the plant concentration at harvest, the annualplant yield, and the soil concentrations at harvest. The calculation is represented by thefollowing equation.

Cli(tQ) = C,,(t)[(C,, - C',(t) Yý)/ C,,] (3)

where

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Csi(t+) = surface-soil concentration for radionuclide i after correction for harvest removalat time t (Bq/m2)

C51(L) = surface-soil concentration for radionuclide i before correction for harvestremoval at time t (Bq/m2)

Cci(t) = crop-c concentration derived from soil uptake for radionuclide i at time ofharvest (Bq/kg)

Y, = annual yield of crop c (kg/m2).

The term in brackets represents the average fraction of the contaminant in soil to that incrops over the year; this term is used rather than a simple subtraction of amount harvestedbecause, for short-lived radionuclides, the amount harvested over the year may actuallybe larger than the amount remaining in the soil at the end of the year. Note that thisformulation assumes that harvested materials are removed from the system, and notreturned (for instance, as manure or fertilizer).

Under natural conditions, the rate of soil removal by erosion generally is in approximateequilibrium with the rate of soil development from soil forming processes, and underthese conditions, soil depth is relatively constant (Troeh et al. 1980). Human activitiestend to accelerate the rate of soil removal. The removal of surface soil by erosion wouldresult in the loss of radionuclides attached to the soil particles. Most models ignore thisloss, assuming that soils lost from one location would re-accumulate in another. Themodel used for analyses for the proposed repository at Yucca Mountain, Nevada, whichis designed to evaluate conditions over periods of thousands of years, does include awind-erosion term (Wu 2003). The rate of radionuclide removal from surface soils isquantified in that model using a first-order surface soil erosion rate constant, defined interms of the surface soil erosion rate as

A,= El/p, d, (4)

where= the surface soil erosion rate constant (yr'1)

E = the soil surface erosion rate (kg m"2 yr")d, = depth of soil rooting zone (m)

Ps = surface soil bulk density (kg/m3).

Application of the soil model for the case of continuous irrigation with a contaminatedsource results in the estimated concentration of contaminant in soil increasing with timeuntil it reaches equilibrium. At equilibrium, the input rate equals the combined loss ratesthrough decay, leaching, erosion, and harvest. For mobile radionuclides (those withsmall Kd values), the equilibrium can occur within a few years of the initiation ofirrigation, as illustrated in Figure 1. However, for radionuclides with a large Kd, and asmall soil-to-plant uptake (which minimizes harvest loss), the approach to equilibriumsoil concentration can take hundreds to thousands of years (Figure 1). This leads to therequirement for an additional input-the length of time irrigation is assumed to occurbefore the exposure of the individual under consideration. Most codes, such as GENII,MEPAS, or RESRAD, can be set to assume that the irrigation begins either at the timethe exposure scenario begins, or at some time before that to allow for buildup. Once

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irrigation starts, it is assumed to continue at a constant annual rate (there may be seasonalstarts and stops).

504535 40 "High Kd

=035- "*d30

C25-

0202-15" LowKd0_ 10

0 20 40 60 80 100Time Since Irrigation Started, Years

Figure 1. Idealized Behavior of Radionuclide Concentration in Soil Following LongPeriods of Irrigation for Mobile and Immobile Radionuclides

An example application of an irrigation model is provided by the Hanford Site SystemAssessment Capability (SAC) (Bryce et al. 2002). In the most recent modelingapplication using the SAC, irrigation is assumed to start at all locations simultaneously atsome fixed time following Hanford Site closure and continue indefinitely (Bryce et al.2002). A slightly different example is provided by the analyses performed for the YuccaMountain Project. In these analyses, the start time for the irrigation scenario wasgenerated randomly, followed by a random period of time before exposure. Because theYucca Mountain analyses use pre-calculated Biosphere Dose Conversion Factors, theanalyses fit a set of factors generated for different periods of time following initiation ofirrigation to an exponential curve. The analyses also used the curve fit as a simple meansof evaluating the impacts of long periods of irrigation (Wasiolek 2001). Both the Hanfordand Yucca Mountain analyses are projected for very long times into the future (up to

* 10,000 years). Continuous irrigation using groundwater has taken place for much shorterperiods (about 100 years) in the United States. Increasing salinity in soils is a long-termproblem. It is still an open question whether irrigation can continue for millennia.

Some radionuclides may be released from soil to air as gasses. This mechanism is only ofconcern for radionuclides that are gases, produce gaseous progeny, or form gaseouscompounds (e.g., 222Rn and 14C). This removal mechanism is not usually included in

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most general models. It would result in non-equilibrium concentrations of decay progenyin the case of radon emanation. The RESRAD model includes estimation of radon fluxesfrom soils; the result is not used to deplete the soil of radon progeny.

Most of the models described above deal with tillage or plowing in an implicit manner.The surface layer of soil is assumed to be uniformly mixed with the radionuclide.Cultivated soil is assumed to be ploughed, if not every year then every few years. Thisgeneral concept is also used for residential soils not used for raising crops. Anothermechanism for mixing of surface soils is bioturbation, the disturbance of soil layers bybiological activity. These biotic pathways include transport of contaminants throughvarious soil horizons by plant root systems and by burrowing insects and small mammals.Bishop (1989) noted that the maximum amount of soil moved by earthworms inestablished pastures is in the order of 10 kg/m2/y (equivalent to a soil depth of around 7

umn). In soil science, bioturbation is modeled as a diffusive process, adopted to avoidquantifying the numerous types of mixing resulting from flora and fauna. The diffusioncoefficient describing bioturbation can be determined by fitting results to verticaldistributions of natural tracers, radionuclides from fallout, or introduced particles. Amore mechanistic approach was developed by McKenzie et al. (1986), which explicitlyevaluated the transport of soil from various subsurface layers to the surface soil. TheMcKenzie et al. (1986) approach provides the amount of soil moved, and the fractionfrom various depth layers, per year. This is summarized in Table 1 for variousenvironmental conditions, which indicates that even in relatively undisturbed soilswithout plowing, mixing will occur to substantial depths relatively rapidly.

Table 1. Biotic Transport Quantities (adapted from McKenzie et al. 1986)

Volume Transported to the Surface (mi3 m"2 yr')Environmental Condition

Depth of Soil Layer Arid Humid Agricultural

< 0.15 m 9.41E-4 7.48E-4 7.48E-40.15 - 0.5 m 7.62E-4 6.73E-4 6.73E-40.5 - 1.0 m 1.79E-4 7.18E-5 7.18E-51.0 - 1.5 mi 1.88E-5 4.49E-6 4.49E-61.5 - 2.0 m 7.53E-6 3.74E-8 3.74E-8> 2.0 m 1.88E-6 3.74E-8 3.74E-8

3.2 Modeling Resuspension

Resuspension relates the concentration of contaminants in air above a contaminatedsurface resulting from losses from that surface to the air. Resuspension occurs whenwind exerts a force on surface material or when there is a mechanical action that disruptsthe surface (such as agricultural operations or vehicular movement). Modelingresuspension is not easy because of the large number of potential variables. The mainones include:

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* Particle size, shape and adherence;* Wind speed,* Surface type and cover (measured in terms of "roughness");* Time since deposition;* Intensity of the mechanical action.

To quantify the suspension of contaminated soil/sediment into the atmosphere, three mainapproaches have been usually used: - resuspension rates, resuspension factors, or massloading. The resuspension rate is a measure of the fraction of material entering the airper unit time, and has units of s-. The resuspension factor relates the radionuclideconcentration in soil (per unit area) to the concentration in air and has units of mr. Theresuspension rate is a mass of soil, assumed to have a contaminant concentration directlyrelated to the source soil, suspended in a volume of air. The mass loading approachessentially assumes that the amount of radioactive contamination in the air is associatedwith dust having the same concentration as local soil. The resuspension factor for foodcrop and animal product pathways is representative of conditions on farmland, whichmay be different from the resuspension factor for the inhalation exposure pathway.Farmland would be expected to be tilled and have soil generally looser than soil for thegeneral residential exposure situations.

The use of resuspension rate approach is based on the transfer of contamination from thesoil surface into the air. The resuspension rate is defined to be the ratio between particleflow density and soil contamination, thus, the resuspension rate dimension is a reciprocaltime. This method provides a rate constant for loss from the surface soil; an additionalmodel is required to convert this into an air concentration.

The use of resuspension factors in the dose assessments is based on the assumption thatthe particulate matter in the air has the same activity as the soil at the location. Theresuspension factor is defined to be the ratio between air concentration and soilcontamination, thus, the resuspension factor dimension is a reciprocal length. This isrepresented mathematically as follows.

Ca = RF C, (5)

where Ca = air concentration of radionuclide (Bq/m3)RF = resuspension factor (mnf)CS -= average surface soil concentration (Bq/m2).

Resuspension factors have been found to cover a wide range (10.10 to l10- m'1),depending on many factors, although this variability is based on short sampling times andlong-term variability will be less (Garger, Hoffman, and Thiessen 1997). This approachallows a convenient method of expressing the observed relationship between surface andair contamination, but with a number of limitations. There is an implicit assumption thatthe air concentrations are a result of the local surface contamination, however, the airconcentration usually includes resuspended materials from upwind sources, which maybe contaminated at different levels. In addition, the time scale of the measurements(generally short) may not correspond to long-term average needed for annual-averagedmodels. A decrease in the amount of resuspension has been observed with time, caused

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by weathering, erosion, or migration of materials down into the soil. This timedependence has been considered by modelers (e.g., Anspaugh et al 1975; CEC 1995;Linsley 1978) and experimentalists (e.g., Garland et al. 1982). Many authors havedescribed the time dependence using exponential factors, but Garland (1979) found thatan inverse power function fitted wind tunnel data over a period of several months. Arecent review (Walsh 2002) recommends the Garland approach modified to account forlong-term resuspension, given as

RF = [1.2 x l0-6 t-' + 10-9]e' (6)

where 1.2x10-6 = resuspension factor at the time of initial deposition to soil (m-)t = time after initial deposition of material to soil (d)10 = resuspension factor after a long time (m').

The second term in this equation is added based on the assumption that no furthermeasurable decrease in the resuspension factor process occurs after the longest period forwhich there are data available. In the implementation, an assumption is made that onlythe top thin layer of soil is available for resuspension.

The atmospheric mass loading of soil in air approach equates the radionuclideconcentrations in soil to measured or estimated dust levels in air. Mass loading is used toestimate the resuspension factor as follows:

RF= S (7)

where S = mass loading of soil in air (g/m 3),p, = surface soil density (g/m3),d, = thickness of surface soil layer (cm).

Different resuspension models may be recommended for different contexts (Garger,Hoffiman, and Thiessen 1997). The resuspension rate and resuspension factor models maybe betters suited for analyses following single acute depositions. The simple mass-loading model may be better suited, and more stable, for analyses of long-term soilaccumulation, because it is not based on measurements from single deposition events, asis the resuspension factor approach.

3.3 Modeling Radionuclides in Vegetation

Radioactive contamination of vegetation may be a major contributor to human healthimpacts through both direct ingestion and/or indirect transfers via the food chain. Plantsmay become contaminated by various processes including direct deposition, externalcontamination with local contaminated soil via resuspension or splashup, incorporationinto edible parts of contamination deposited on leaves or stems, and uptake via the roots.The basic components to be modeled are illustrated in Figure 2.

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Figure 2. A Basic Model for Contamination of Plants

3.3.1 Modeling Foliar Interception and RetentionDeposition of airborne contamination is largely a function of particle size and.precipitation. Radionuclides can be present in air as single atoms, as chemicalcompounds, or as contaminants attached to dust or other aerosol particles. As a generalrule, one can expect the deposition of radionuclides to be governed by the physicalprocesses involving aerosol-land surface interactions. The rate of radionuclide depositionfrom the atmosphere to the ground is proportional to concentrations in air, although theactual deposition is also affected in complex ways by particle characteristics, receivingsurface properties, and weather conditions.

Airborne radionuclides can be deposited onto soil or vegetation by "dry" mechanismssuch as gravitational settling, surface impaction and electrostatic attraction and by "wet"mechanisms such as washout or rainout during precipitation events (Kinnersley andScott, 2001). As a general rule, particles with diameters larger than about 20 pm anddensities higher than I g cnf- are affected more by gravity than atmospheric turbulence(Whicker and 'Schultz, 1982). Deposition of these larger particles can be predicted usingStoke's law to evaluate a gravitational settling velocity. Smaller particles that areaffected more by atmospheric turbulence than gravity can travel great distances beforedepositing. Deposition of these smaller particles is affected not only by particle size anddensity, atmospheric turbulence and precipitation, but also by the nature of the plantsurfaces that the particles land on.

Contaminants may also be deposited through human activities such as irrigation. Thetypical irrigation cycle can be from one to many hours. Once plant surfaces are initiallywetted, subsequent water flows to soil as rnm-off. The question is how much of aparticular element can be entrained, adsorbed, and/or absorbed during an irrigation event.Nair (Nair et al. 1996) found in rainfall simulation experiments that larger particlestended to adhere to the surface of vegetation and that subsequent rainfall was ineffectivein removing what had interacted with the leaf surface during an initial rain event. Foranions, like various forms of iodine, the radionuclides followed the water. But, if drying

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occurred, then subsequent rainfall was ineffective in removing iodine from the leafsurface. This would likely be true for chemical forms having low to intermediateabsorption rates and for elements with low adsorption rates, or with high absorption rateswith limited leaf-surface binding sites. It may not be true for elements with exceptionallyhigh bioavailability, as would possibly be the case for technetium and nickel.

Many models, e.g., RESRAD, MEPAS, and CAP88-PC (Chaki and Parks 2000), use aconstant dry interception fraction of 0.25. The MILDOS code uses 0.2. The IAEA(2001) Safety Standard model uses a mass interception factor, defined as the fraction ofdeposited activity intercepted by the edible portion of vegetation per unit mass (or massinterception factor, m2/kg), which is set equal to 3 m2/kg (dry weight) for forages and 0.3m2 /kg (wet weight) for foods consumed directly by humans.

For a specific deposition event, the interception fraction is higher the more developed theplant canopy, which is accompanied with an increasing contact surface between thefalling rain and the plant surface. There are two principal approaches to parameterize theplant development, the standing biomass per unit soil surface area and the leaf area perunit soil surface area (leaf area index). The advantage of the standing biomass is that it iseasy to determine. The sample has simply to be taken from a known area of ground,dried, and the dry mass has to be determined. However, the biomass does not reallyrepresent the size of the interface rain/plant, which is the case for the leaf index. Whereasin the first period of growth, in general a good relationship between biomass and leaf areacan be observed, the correlation fades away towards the end of growth. Then, thebiomass still increases due to growth of storage organs as seeds or tubers while the leafarea already decreases substantially due to dying off of the foliage. The disadvantage ofthe leaf area index is that quantification is a complicated determination that requiresspecific optical devices which are usually not available. However, a wide variety of dataon leaf area indices for terrestrial systems has recently become available from NASAsatellite imagery as a result of their involvement in carbon cycle and climate research,which may simplify this problem.

Models that at least account for biomass have been proposed. An empirical relationshipbetween biomass and interception fraction from atmospheric dry deposition wasoriginally suggested by Chamberlain (1967). This model has been expanded by Pinder etal. (1988) for grasses and other species. The following form of the equation is suggested:

rd, = 1-- e-A 4o f, (8)

where

rdC = interception fraction for atmospheric dry deposition to crop type c (dimensionless)

A = empirical constant, about 2.9 for grasses, leafy vegetables, and grains, and 3.6 forfruits and other vegetables

Y, = standing biomass of the growing vegetation for crop type c (kg wet weight/m2)

f, = dry-to-wet weight biomass ratio for crop type c (kg dry weight per kg wet weight).

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The dry-to-wet ratio is required because the Pinder formulations are given in terms of drybiomass. This formulation results in the need to define the growing biomass as well asthe harvested yield.

The interception of material in irrigation water is not well studied. Thus, a default valueof 0.25 is often used for all materials deposited on all plant types by irrigation. Theresults described above for irrigation are probably applicable. An empirical equation forthe interception fraction, r., used in the Yucca Mountain model ERMYN (Wu 2003)derived from data from Hoffman et al. (1989), which is based on the results ofexperiments with 7Be and 1311, is expressed as

=IA' 1 (9)

whererw, = interception fraction of irrigation water for crop type c (dimensionless);K1, K2, K3, and K4 = empirical constants (KI is in units of (kg/m2)"rK (mm)"rK

(cm/hr)'K4,and K2, K3 and K 4 are dimensionless);Yc = standing biomass of crop type c (kg dry weight/m2);IAc = amount of irrigation per application event for crop type c (mm);I = irrigation intensity (value in units of cm/hr).

Because this is a regression equation from experimental data, values for the inputparameters must be used in the units specified above. The empirical constants in thisequation were developed based on given parameter units for standing biomass, irrigationamount, and irrigation intensity, depend on the plant type and contaminant form. Therecommended values (Hoffman et al. 1989) are:

K, = 2.29 for beryllium (Be+); Ki = 1.54 for iodine (I-)K2 = 0.695 for beryllium (Be+); K2 = 0.697 for iodine (I-)K3 = -0.29 for beryllium (Be+); K3 = -0.909 for iodine (I-)K4 = -0.341 for beryllium (Be+); K4= -0.049 for iodine (IL)

The interception fraction for wet deposition can be also evaluated as a function of therainfall rate and standing biomass based on experimental observations of Prohl andHoffman (1993). The data by Prohl and Hoffman were used to develop relationships foranions, cations, and for insoluble particles for the GENII Version 2 application (Napier etal. 2002). For anions, such as iodide and sulfate, the interception fraction is evaluated asfollows:

r, =2.3 Y, f, R -092 (10)

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where

rw, = interception fraction from wet deposition to crop type c (dimensionless)

Y, = standing biomass of the growing vegetation for crop type c (kg wet weight/m2)

f, = dry-to-wet weight biomass ratio for crop type c (kg dry weight per kg wet weight)-

R = rainfall rate (mm/d),

and the numbers are regression parameters.

For cations and particles, Prohl and Hoffman (1993) presented the following relationshipbased on experiments using microspheres (essentially the same data as Hoffman etal. 1989):

r,• =2.95 Yý f, R-'191 (11)

where terms are as previously defined. In both equations, the numerical term is a fittingparameter with units of m2/kg. A potential disadvantage of this approach based onbiomass is that it is intended primarily for leafy-type produce; its indiscriminate usecould result in the more massive fruits such as apples, melons, tomatoes interceptingmore irrigation water than the less massive but more expansive (e.g., surface area, leafindex) lettuce or spinach plant. The selection of which biomass to use is important.

Beyeler et al. (1999) also evaluated the results of Prohl and Hoffman (1993). Beyeler etal. (1999) decided that it was difficult to assign anion/cation status without site-specificknowledge of groundwater chemistry and elected to use a single range of interceptionfractions. They recommend a central value of about 0.35 within a uniform range of from0.1 to 0.6 for all crop types.

The ECOSYS code uses a formulation based on Leaf Area Index (LAI) and the water-storage capacity of plant leaves. Buildup of the water film on the leaves during wettingevents, the total amount of water deposited, and the radionuclide's ability to be fixed onthe leaf (another interpretation of the data of Prohl and Hoffrnan) are considered. In thisformulation, the wet interception fraction is estimated as

= LAI, Si / R [I - exp( 31 2) R]

where LAI is the leaf area index of plant i, Si is the retention coefficient for plant type i(mm), and R is the amount of rainfall (mm).

If this equation results in an interception fraction greater than one, the value is set to one.Muller and Prohl (1993) provide estimates for the retention coefficient. As for themodified Chamberlain model described above, these vary for grasses, cereals, and corn,and for other plants.

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Measured interception fractions for a limited number of radionuclides are becomingavailable for Asian food types not usually included in American risk assessments such asrice, radish, soybeans, and Chinese cabbage (Choi et al. 1998; 2001; 2002; Lim et al.2001).

Many investigators have commented that the plant biomass and the interception fractionare directly correlated, and a more stable estimate of interception may be obtained usingthe quotient r/Yas a single variable (IAEA 1982). The interception fraction r rapidlyincreases with the increasing plant biomass during the first half of the growing period andthereafter it is almost constant or even decreases (Vandecasteele et al. 2001). In contrastto r, the biomass-normalized interception fraction r/Y decreases as plants grow from thebeginning. The r/Y values at the early growth stages are conspicuously high, becausemost of the young leaves are directly exposed to deposition with comparatively lowmass-to-area ratios. Accordingly, the rlY is a useful parameter in estimating the initialplant concentration after an accidental deposition but it is unnecessary during routineconditions for which such an initial concentration cannot, and need not, be calculated. Forassessing continuous deposition, instead, an averaged value of r/Y can be used, or simplya lumped parameter that is a ratio between the deposition rate and the steady stateconcentration per gram of plant material.

Kinnersley and Scott (2001) suggest that the complex three-dimensional structure of cropcanopies is such that no purely mechanistic model yet exists which can predict theturbulent transport of airborne particulate contamination into, and onto, them. Whiletheoretical understanding of the processes involved in particle dry deposition hasadvanced considerably in the last decade or so, a gap remains between theory and thepractical prediction of particle deposition, particularly under non-laboratory conditions.Bonka and Horn (1983) produced a thorough review of the mechanisms governingdeposition of aerosols to vegetation, from which it can be deduced that analyticaldeposition models are currently limited by uncertainty in the aerosol collection efficiencyterm. This term must be determined experimentally for canopies consisting of anythingbut regular arrays of the simplest geometrical shapes. The review describes the model ofSlinn (1982), in which the maximum possible deposition velocity (assuming the canopyto be a perfect sink) is attenuated by a term, ultimately empirical in derivation, allowingfor the particle collection efficiency of the canopy. Data requirements for this modelinclude a "vegetative hair fraction", a "small vegetative diameter" and a "large vegetativediameter". Another well-known model for deposition to vegetation is that of Sehmel andHodgson (1978). This model takes into account the underlying physics of particleinterception in some detail, including amongst its input data requirements details of theturbulence, boundary layer and surface layer resistances, but again it is ultimatelydependent on empirical terms determined in a wind tunnel. A third approach has beenadopted in the Imperial College (IC) model. This is based on multiple regression analysesof a considerable database of deposition values generated under wind tunnel conditions.The regressions use the three parameters - particle size, turbulence levels and canopy size- which theory suggests should dominate control of the deposition process. These threemodels give broadly similar results, which are compatible with the few reliable fieldmeasurements of deposition to crop canopies which currently exist. Despite data

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requirements which can become complex, current deposition models ultimately retain adependence upon experimentally determined factors. However, there is a lack of reliabledata of this sort. This is particularly so for the interception of particles less than 40 rnm indiameter (Chamberlain & Garland, 1991) and for larger canopies such as those of fruitcrops.

For prolonged deposition to vegetation (as occurred after weapons testing), Chamberlainintroduced the concept of normalized specific activity (NSA), defined as the ratio ofradionuclide concentration in foliage (Bq/kg dry weight) to the rate of deposition toground (Bq/m2/d). A value of 40 m d/kg was considered appropriate for herbage in goodgrowing conditions; much lower values have been found for grains (Simmonds andLinsley 1982). NSA values are much higher for poor growing conditions and some typesof native vegetation (Martin and Coughtrey 1982).

3.3.2 Modeling Contamination of Plant Surfaces with Soil

Plants may acquire surficial contamination in the form of dust or soil particles; the sourceis the soil in which the plant is growing. Certain types of plants, such as leeks, arefrequently associated with soil carryover.

Contaminants sorbed to soil particles may be resuspended by wind, human, animal,mechanical disturbances and redeposit locally on the plant surfaces. Such redepositionoccurs when soil is displaced by wind (Anspaugh et aL 1975), animal disturbance(Hinton et al. 1995; Sumerling et al. 1984) or mechanical disturbance (Milham et al.1976; Pinder and McLeod 1988; 1989). Another physical mechanism for contaminationof plant surfaces with soil particles is raindrop splash (Dreicer et aL 1984). In the 30-kmnzone around Chernobyl, rainsplash of contaminated soil contributed significantly tooverall plant contamination, with 60 to 70% of the activity concentrated on the basalparts (10 to 20 cm above the soil surface) of cereals and perennial grasses (Kryshev1992).

Plant contamination may also be caused by mechanical disturbances that raise dust, suchas occur during weeding and harvesting. Adriano et al. (1982) and Pinder and McLeod(1988; 1989) have reported this to be a major source of plutonium contamination in graingrown on contaminated soils. Plants may acquire surficial contamination in the form ofdust or splashed-up soil; the source is the soil in which the plant is growing. Soil-to-Vlanttransfer values via the roots for some actinide elements are on the order of 10.6 to 10°

(Bq/kg dry crop)/(Bq/kg dry soil). This very low transfer from soil via the roots meansthat the contamination of a crop due to contaminated soil resuspended or splashed on theleaf surface may easily dominate the plant physiological uptake. A contamination of only1 mg soil per kg fresh weight crop can cause an apparent transfer value of 10'5 forsplash-up and resuspension transfer factors. Thus this pathway may be of importance insome analyses of radionuclides that are biologically discriminated against.

The simplest approach is to define a soil adhesion quantity for each type of crop plant.This combines the effects of interception of locally-resuspended material and thesplashing of rain and irrigation water droplets, and allows averaging of the effects of

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various factors such as height and type of plant canopy, wind, rain, and soil type. Thisapproach is suggested by IAEA (1994) and used by Kennedy and Strenge (1992). Thelater IAEA model (LAEA 2001) combines this mechanism with the soil-to-plant transferinto a single transfer parameter (see below) - to account implicitly for soil adhesion, aminimum value of 0.1 is assigned to the soil-to-plant transfer factor for forages and 0.001to the transfer factor for human foods. However, this approach is not recommended bythe International Union of Radioccologists, who state that "Values of soil adhesiondepend so much on environmental conditions that it is not realistic to propose a defaultvalue. However, an estimate for a low value of soil adhesion is 4 g soil per kg dryvegetation for leaves and grass taken from 40 cm or more above the soil surface. Below40 cm, 10 g per kg might be a reasonable estimate for leaves and grass. Considerablylower values are expected for products such as grain, which are protected by plant partswhich are not consumed. A reasonable upper limit for soil adhesion is 250 g per kg."

The processes of suspension and deposition are parameterized as consisting ofresuspension of contaminated soil followed by local redeposition in some models. This isthe approach taken by the GENII family of codes, as well as the RESRAD family ofcodes, PATHWAY, and ECOSYS. The resuspension factor for food-crop and animal-product pathways is representative of conditions on farmland, which may be differentfrom the resuspension factor for the inhalation exposure pathway. Farmland would beexpected to be tilled and have soil generally looser than soil for the general residentialexposure situatious.

Many models circumvent the necessity of dealing with soil adhesion by specifyingamounts of soil ingested by humans and domestic animals. However, this approach thenneglects the potential for transfer from the plant surface to the internal portions of theplant. It also complicates the modeling of weathering of deposited material off of plantsurfaces, discussed below.

3.3.3 Modeling Weathering from Plant Surfaces

Material deposited on vegetation may be lost from the plant through a variety of differentprocesses in addition to radioactive decay, such as removal by wind or rain,volatilization, plant senescence (leaf drop), and others. Wind is thought to dominate thefield-loss process for particulate contamination, with the contribution of rainfall beingsignificantly less (Kinnersley and Scott 2001). In fast-growing vegetation, a decrease inconcentration may be seen as a result of dilution by new vegetative material, and ascribedto a removal process.

Losses from plant Furfaces are frequently represented by a weathering rate constant, X,.Several processes may be involved. Some measurements of this process have implicitlyincluded plant growth. Volatile materials may evaporate back into the air. Ifprecipitation occurs, or there is heavy irrigation, physical washoff processes may bepresent. Because of the combination of potential processes, this phenomenon issometimes calledfield loss.

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The lack of detailed measurements of contamination loss over time has led to the use of asimple exponential model for weathering (Chadwick and Chamberlain 1970). TheNRC's Regulatory Guide 1.109 (NRC 1977) uses a weathering half-time of 14 days. Thedefault-rate constant in GENII, and in PATHWAY for all radionuclides other thaniodines, is based on a weathering half time of 14 days. The values in IAEA Safety Series57 (IAEA 1982) are 15 days for particulates and 10 days for iodines; that used in theIAEA Safety Reports Series 19 (IAEA 2001) is 14 days. The value in RESRAD (20yearf ) is equivalent to a half-time of about 12.6 days. ECOSYS uses a half-time of 25days; the ECOSYS model attempts to account also for dilution via plant growth inpasture grasses. When combined, the ECOSYS results are in the range of 10 to 16 days.Statistical analysis of data from literature shows a log-normal distribution for half-livesand gives a default value (geometric mean) for estimating the removal of radionuclidesfrom vegetation based on a half-life of 17 days'.

Kinnersly and Scott (2001) report that the dominant weathering/field loss processes arethe resuspension of particles into the air by atmospheric turbulence, and loss throughbiological processes such as leaf-loss and cuticular wax-shedding. In recent years muchhigher resolution studies have been carried out (Ertel, Voigt, and Paretzke 1989; Fraley,Chaveg, and Markham 1993; Kirchner 1994; Kinnersley et al. 1996), and it is now clearthat field loss is best described in terms of rapid and slow loss components. The slow losscomponent can be so slow as to be negligible, such that it can be treated as a fixedfraction of contamination that will not be lost over time, yielding a model (not includingradioactive decay) of the form (Kinnersley et al. 1997):

A, =A, + (Ao + A,) eA' (13)

where A0 is the initial level of contamination, A, is the level of contamination at time t, Ar

is the fraction of the original deposit not subject to loss, and X is the time constant for thefraction of initial deposit subject to exponential loss. This particular empirical.formulation may also represent the fraction of material that is absorbed and/ortranslocated by the plant versus that which remains on the surface and is available forweathering (see the next section).

The simple one-exponential approach or the NSA are the easiest to implement insituations with continuing deposition; the Kinnersly et al. (1997) formulation is bettersuited to single deposition events. Similarly, models such as that in ECOSYS thatattempt to account for dilution of contamination in growing crops are better suited forsingle depositions; a long-term average field loss is better suited for continuousdepositions where the intended use is determination of ultimate concentration at harvest.

E. Leclerc, unpublished data prepared for the IAEA's Environmental Modeling for RiskAssessment coordinated research program.

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3.3A Modeling Foliar Absorption and TranslocationAbove-ground parts of plants (leaves, blossoms, fruits, and branches) can directly absorbwater and minerals as well as soluble radionuclides. This process is called foliarabsorption. Some radionuclides, once absorbed into the body of the plant, will remain ator very near the location from which they were absorbed. Other radionuclides may betransported via the phloem of the plant to other plant structures, organs, or fruits. Thisprocess is referred to as translocation.

The processes of absorption and translocation are dependent on the physical and chemicalforms of the contaminating radionuclides. Materials incorporated in relatively largeparticles are not easily absorbed, and so are unlikely to be affected by translocation.Contaminants in ionic form are more likely to be absorbed.

The foliar absorption of 13 7Cs that was sorbed to resuspended soil particles was quantifiedat two sites contaminated by the Chernobyl accident. Although measurable, this foliarabsorption was considered inconsequential relative to other plant contaminationpathways, which led to the recommendation that it not be considered a critical pathway inroutine radionuclide transport models (Hinton et al. 1996).

Once absorbed, the processes of weathering from the plant are changed; transport bywind and rain are much reduced, although physical losses such as leaf drop will remain.

The translocation factor indicates the fraction of total deposition to plant surfaces that isincorporated into edible parts of the plant. This mechanism incorporates two processes -Initial uptake by the leaf and subsequent translocation throughout the plant is dependenton the cuticular properties of the plant, its wetability, strength of absorption onto thecuticle, the radionuclide in question, its solubility, size, speciation, and the time betweencontamination and harvest. The importance of uptake and translocation will bedependent on what part of the crop is to be eaten and what species of radionuclide hasbeen intercepted.

In many models, e.g., GENII (Napier et al. 1988; 2002), RESRAD (Yu et al. 2002), andCAP88-PC (Cahki and Parks 2000), the translocation factor is defined to be the fractionof total deposition to plant surfaces that is incorporated into edible parts of the plant. Avalue of I is assumed for leafy vegetables and forage crops where the whole above-ground portion of thp plant is eaten, and 0.1 for all other vegetation. The value of 0.1 isassigned on the basis of very little information and is assumed to be an upper bound.

The IAEA (1994) provides two alternative definitions for the translocation factor, as

T = activity concentration in the edible parts at harvest (Bq / kg) (14)activity retained on I m2 of foliage at the time of deposition (Bq m 2 foliage)

Ta activity concentration in the edible parts of I m2 at harvest (Bq M m2 )activity retained on I m2 of foliage at the time of deposition (Bq / m2 foliage) (15)

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Note that each of these formulations has different units. IAEA (1994) recommends usingthe first approach, normalized to the yield in kilograms, as having less measurementvariability. A later IAEA safety series model (IAEA 2001) includes translocationimplicitly in the definition of the foliar interception factor used in that (intentionallyconservative) model.

The relative rates of absorption of chemical elements from cuticular structures aredifficult to quantify. First, absorptive capacity differs from root processes in that thereare apoplastic/tricome routes of entry into the foliar interior, which are lessdiscriminating (with respect to species discrimination) than for roots. Secondly, thetransfer of individual elements is very dependent on chemical form and stability. Forexample, free cations and stable anions may be readily absorbed, while reactive orunstable species may not be absorbed, or absorbed to a lesser extent. The transferprocess is basically a molecular level interfacial film flow. Thus, only molecular-sizedmaterials can move. Finally, refractory particles, i.e., oxides, carbonates, and hydroxides,of any size will not likely be transferred to the tissue interior unless a significant rate ofsolubilization is occurring, e.g., complexation agents in irrigation waters: These particlescan and will, however, be entrained within the complex cuticular structures associatedwith most surface foliage.

There has been much research into agricultural chemicals applied to foliage, some ofwhich has been performed with radioactive tracers (e.g. starting with Tukey and Wittwer1956; Tukey et al. 1956). Potassium and sodium were found to be rapidly absorbed andhighly mobile. Phosphorus, sulphur, and chlorine were absorbed at a slower rate, butwere also mobile and were transported at a rapid rate. Manganese, zinc, copper, andmolybdenum were found to be slightly mobile. Calcium, strontium, barium, iron, andmagnesium were readily absorbed but did not move out of the leaf to which they wereapplied. McFarlane and Berry (1974) reported that the rate at which cations penetrate thecuticle is inversely related to the radius of the hydrated ion, which occurs in the orderCs>Rb>K>Na>Li>Mg>Sr>Ca. Aarkrog (1975) reported that different radionuclides weretranslocated to a varying rate and degree in the order:Zn>Fe>Cs>Co>Mn>Na>Cd>Sb>Cr>Sr--Ce-Ru>Ba-Hg-Pb. This pattern may varywith the chemical state of the element and the presence of complexing agents. Thesolution sprayed on the crops had a pH of 1.6 and solubilities would be diferent fromthose in actual rainfall. Also, only the activity rather than the species of the radionuclidesapplied is known so the results may not be representative of a realistic situation. Bukovacand Wittwer (1957) suggested mobility in the order:Rb>Na>K>P>CI>S>Zn>Cu>Mn>Fe>Mo>Ca>Sr>Ba. Distribution of Rb, Na, K, Cl, P,S and possibly Zn appeared proportional to the metabolic activity of various plant tissues.Foliar absorption is time dependent, i.e., the longer a contaminant is retained on a leaf thegreater is its absorption. Higher absorption rates have been documented in younger leaves(Hull et al., 1975), and significant percentages of the total cesium intercepted by foliagecan be absorbed within a few days (Carini and Bengtsson 2001).

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While little data exist to demonstrate or quantify foliar absorption processes andparticularly rates of transfer, a list of probable behaviors for priority radionuclides isgiven in Robertson et al. 2003. This categorization is based on known behavior insoil/plant systems and likely chemical speciation in groundwaters. This listing is basedon many assumptions. Although much data exist for nickel and technetium, little isknown of the foliar routes of entry for any of these elements. The intermediate categoryis questionable, but the limited data available would indicate that there is a need for studyof particularly uranium, iodine, and americium. In the proposed Low Transfer category,little is known about the behavior of beryllium and niobium.

In the RODOS code (MUller, Gering, and Pr6h1 2004), the translocation process isquantified by a-translocation factor defined as the fraction of the activity deposited on thefoliage being transferred to the edible parts of the plants until harvest. It is dependent onthe element, the plant type, and the time between deposition and harvest. In thisformulation, only a single deposition is assumed, so the time. from deposition to harvest isimportant. The elements are grouped into two sets, mobile and immobile; the mostreliable data for translocation are available for Cs and Sr. Therefore, as long as no otherquantitative data are available, the translocation of mobile elements assumes that theybehave like Cs, except for Mn, for which translocation factors are assumed to be lower bya factor 0.6 than for Cs. Strontium is considered to be representative for immobileelements. The RODOS assignments are also shown in Table 2; it can be seen that there issome coincidence of interpretation with Robertson et al. (e.g., Tc, Pu, Sr), and somedisagreement (e.g., Np). Note that the RODOS implementation tends to focus onradionuclides with shorter half-lives.

Table 2. Probable Bioavailability/Cuticular Transfer of Selected Elements. This refers tofoliar uptake, and not specifically phloem mobility and redistribution (adaptedfrom Robertson et al. 2003 and Miller, Gering, and Pr6hI 2004)

High Transfer Intermediate Transfer Low TransferNp Am PuTc Cs ThNi U Be

Cm NbI Sr

SeAgSn

Mobile ImmobileCo, Cs, I, Mn, Mo, Na, Ag, Am, Ba,.Ce, Cm, La,Rb, Sb, Tc, Te Nb, Nd, Np, Pr, Pu, Rh,

Ru, Sr, Y, Zr

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There is unfortunately no in-depth compilation of the rates of retention nor uptake ofradionuclides into the plant via foliar contamination. Kinnersley and Scott (2001)undertook model parameterization for wet and dry deposition processes to plant surfaces.This theoretical parameterization is reasonable, but there is a nearly total lack ofsupporting data. The IAEA (1994) provides a limited number of rate constants for a fewelements for a few crop types.

Translocation and weathering are, in some senses, competing processes for transport ofmaterial off of plant surfaces. This is often overlooked in model formulation: materialtranslocated may be affected by the weathering terms if not separately.considered. TheERB2A model from the IAEA BIOMASS program (LAEA 2003) and the YuccaMountain ERMYN model (Wu 2003) are examples that specifically separate these twoeffects. This is done by developing a fraction of the surface contamination to whichweathering processes apply - the difference between the total and the translocationfraction. These models do this by adding additional parameters in the following fashion:

C = - F..)e-Ay FP3 + F F, 2¥ F•,•,,.1 (16)

Cc, img= radionuclide concentration in the edible part of the crop due to irrigation waterdeposited on crop leaf surface (Bq/kg fresh weight of crop)

r, = fraction of radionuclides in spray irrigation water initially deposited on standingbiomass (dimensionless)

S = annual irrigation deposition (Bq m2);Ftr,,, = fraction of absorbed activity translocated to edible portions of the plant by the

time of harvest (translocation fraction)Fabs = fraction of intercepted radionuclide initially deposited onto plant surfaces

absorbed from external surfaces into plant tissues (dimensionless)F,2 = fraction of internal contamination in the'edible parts of plants at harvest that is

retained after food processing (dimensionless)Fp3 = fraction of external contamination from interception retained on edible parts of

the plant after food processing (dimensionless)AW = removal (weathering) rate for radionuclides deposited on plant surfaces by

irrigation (weathering processes include mechanical weathering, wash-off, andleaf fall) (yr' )

T = interval beiween irrigation and harvest (y)Y, = biomass of the crop (kg fresh weight/mi).

3.3.5 Modeling Soil-to-Plant UptakeTerrestrial plants, as sessile organisms, have adapted to derive essential nutrients fromtheir environments. Plants absorb nutrients through their roots and transport them via thephloem to active portions of the plant. Biotic factors are likely the source of much of thevariability seen in concentrations of contaminants in plants. This variability results fromthe nature of the sessile terrestrial plant and its relationship with its environment: the needto compete and acquire specific nutrient species from soils (or to avoid uptake of

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excesses of potentially toxic materials), the need of individual plant types for specificlevels of individual nutrients, and the use of certain metals in tissues as agents todiscourage herbivory. Thus, the uptake is affected by the following biotic factors:

1. the plant-available concentration in soils within the rhizosphere (the soil zone thatsurrounds and is influenced by the roots of plants), which is governed by soiladsorption processes, chemical solubility, microbial/fungal activity, and stability ofthe chemical complexes

2. the chemical nature and stability of the cation-anion/complex with respect to theplant's capability to metabolically alter and/or absorb the elemental form into theplant

3. a series of plant adaptive/evolutionary processes for survival; this can include, but notbe limited to, protective root processes (exclusion or complexation of an ion fordetoxification, sequestration within the plant to regulate both ion levels and fordetoxification, redox to alter solubility and transport when necessary, organiccomplexation in the case of all but mono-cationic elements, and uptake capacitybeing dependent on metabolic needs).

Under these chemical and biotic constraints, uptake can be expected to vary based onsource term, kinetics of solubilization/speciation, the relative ability of a plant to view anon-nutrient ion as an analogue to a nutrient species, and the relative need by theindividual plant genus/species for specific levels of a particular ion.

Transfer of radionuclides from soils into plants is one of the key mechanisms for long-term contamination of the human food chain. There are several methods that maypotentially be used. The most common is through the use of soil-to-plant transfer factors(or concentration ratios), but others may also be used.

3.3.5.1 Development of Sofi-to-Plant Transfer FactorsResearch on the relationship between radionuclides in soil and in the plants grown in thatsoil began soon after the establishment of the major radionuclide production centers ofthe Manhattan Engineer District. In the early period (through the early 1950s), thepotential for root uptake was recognized (Cline and Porter 1951) but not appreciated; themain concern was foliar deposition (Parker 1955). However, some studies wereoriginally undertaken at the Hanford Site to investigate this pathway (e.g., Seiders, Cline,and Rediske 1955). Hanford's Botany Field Station was established in 1949 and usedColumbia River water from downstream of the reactor outfalls for an irrigation source.Minor increases in gross radioactivity were noted in crops in 1949 and 1950 (Cline andPorter 1951), but individual radionuclides were not identified. However, by 1953, workhad been done at Hanford to determine the concentration factor of several keyradionuclides from soils contaminated elsewhere by fallout (Rediske and Selders 1953a;1953b; Selders, Rediske, and Palmer 1953).

Weapons testing was conducted in the United States and elsewhere and a number ofcountries became concerned about world-wide fallout, particularly of 9°Sr. Results of

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surveys of contaminated biota on contaminated islands in the Pacific Proving Groundswere first published in 1956 (Atomic Energy Commission 1956). Meanwhile, agriculturalscientists in the U.S. and Great Britain began investigations and field studies on thefactors governing uptake of 90Sr and its transport into food chains that led directly tohumans. Concern about uptake through plant roots, as opposed to foliar deposition fromfallout, first arose at Oak Ridge in the mid-1950s following the draining of White OakLake (Auerbach and Reichle 1999; Stannard 1988). Recognition of root uptake ofselected contaminants (primarily related to situations concerning waste disposal in near-surface soils) was wide-spread by the early 1960s; interestingly, however, paperspresented at the first major conference on radioecology (Shultz and Klement 1963)contain data on radionuclide concentrations in plants - but no soil-to-plant concentrationfactors per se. Senate hearings in 1963 focused on radioactive fallout from the weaponstests and food chain contamination (Stannard 1988). The initial attempts to model uptakeof radionuclides by plants were driven by unanticipated contamination events followingatmospheric nuclear weapons testing.

Menzel (1965) presented results of a literature study of various soil science experimentsof soil to plant transfer at the 1965 Hanford Biology Symposium. This is one of the firstuses of the term "concentration factor" with regard to plant/soil ratio. Menzel used thisratio to summarize a large amount of data for several elements (in terms of ppm of dryplant per ppm of dry soil). He was more interested in exploring the effects of radiationon the plants and the mixing of radionuclides in soil by plants than on the actual uptake.(At this same Symposium, Mistry et al. (1965) gave detailed measurements of variousherbs, shrubs, trees, fruits, and tubers grown in areas of high natural background - but didnot provide either concentration ratios or soil measurements from which concentrationratios could be computed.)

The concept of using nuclear explosives for excavation, particularly of a sea-level canalacross the Isthmus of Panama, with the potential for associated fallout, led to severalattempts to predict the potential environmental affect. Martin et al. (1970, as reprinted inKlement (1982)) proposed a series of physics-based, differential equations to describe thetransport of radioactive contaminants resulting from canal-building fallout. This effortincluded estimation of transfer of radionuclides from soil water into plants that wasrelated to the growth rate of the plant and the stable element content of the plant. Noexamples are given for this mechanistic approach; it does not appear to have been toosuccessful.

The document series Prediction of the maximum dosage to man from the fallout ofnuclear devices (Tamplin 1967; Ng and Thompson1966; Burton and Pratt 1968; Ng et al.1968) also came about because of concerns about predicting doses from canal-buildingfallout. This series was intended to provide input to design of nuclear explosives tominimize the impacts of fallout; one segment of the calculation was to develop a "unit-rad" value that related normalized fallout amounts by radionuclide to committed internaldoses. While the "unit-rad" calculation did not directly involve estimation ofradionuclide transfer from soil to plants, a series of appendices to Ng et al. (1968) diddescribe how global average concentrations of stable elements in soils and plants could

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be ratioed to give an estimate of uptake of radionuclides in plants from contaminatedsoils. In this description, Ng used the symbol (Cp/Cs) for what he called the plant-to-soilconcentration ratio. The data in this report became the basis for many later models andcalculations, such as NRC Regulatory Guide 1.109 (1977).

The concept pioneered by Ng has been given many names over the years.* Plant-to-soil concentration ratio - a generic term in ecology, used by Whicker

and Schultz (1982), Peterson (1983), Faw and Shultis (1999), and others;* Soil-to-plant transfer factor - used to describe the direction of radionuclide

migration "from soil into plants", adopted by Martin Frissell and the InternationalUnion of Radioecologists (1989), IAEA (1994), and others;

* Vegetable/soil transfer factor - used in documents related to the RESRAD familyof computer codes (e.g., Yang, Biwer, and Yu 1993; Yu et al. 2001);

* Soil-to-plant concentration ratio - a variant used by NCRP (1984);* Soil-to-plant concentration factor- a variant used by IAEA (1982), Ng, Colsher,

and Thompson (1982);* Concentration factor- a variant used by Soldat in Fletcher and Dotson (1971);

Miller et al. (1980);• Transfer ratio - Stannard (1988);• Concentration ratio - a commonly-used variant (e.g., Cataldo, Wildung, and

Garland 1983; Napier 2003; numerous others);• Plant uptake factor - Eisenbud (1987);• Relative ratio - Dahlman and Van Voris (1976);* Uptake ratio - Grummitt (1976);* Plant bioconcentrationfactor - a definition more usually used for aquatic uptake

processes, but sometimes used with terrestrial vegetation (e.g., Wolterbeek, vander Meer, and Dielemans 2000);

* Plant bioaccumulationfactor - a definition more usually used for aquatic uptakeprocesses, but sometimes used with terrestrial vegetation (e.g., Wolterbeek, vander Meer, and Dielemans 2000)

* Discrimination factor - the name sometimes given to the ratio of the steady stateoutput to input of a compartment model, when applied to the ratio ofconcentration in soil to concentration in plant, also accumulation factor (Peterson1983; Faw and Shultis 1999).

All of these concepts are the same; the notation generally used is B, for the ratio of theconcentration of an element in a plant of interest to the concentration in the source soil.The transfer factor applies to long-term, chronic exposures and is ideally measured atequilibrium. Transfer factors are used in risk assessments to estimate the amount ofradioactivity that could be present in a food crop based on the calculated concentration inthe source soil. By calculating the concentration in the food, the total intake can beestimated and a dose calculated as a result of the annual intake. In terms ofradionuclides, the transfer factor is used to calculate how many becquerels per kilogramof soil are transferred to the edible dry plant product (Bq per kg). Although the conceptis occasionally questioned for some of its underlying assumptions (e.g., Centofani et al.2005), it is the most commonly used approach in radiological environmental assessments.

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The concentration ratio depends on the radionuclide, the soil type (which may includesoil chemistry and concentrations of nutrients and analogues), and the plant type. Thetransfer factors are empirically derived; they are based on measurements made forvarious chemical forms of the radionuclide on selected types of plant in selected soiltypes. Experimental data are not available for all elements for all food types. Frequently,a few measurements on a very limited number of plant types are used to infer a transferfactor for all crops. Often, when no referenceable documents are available, data arederived based on chemical groupings in the periodic table of the elements, such thatchemically similar elements are assigned similar values.

In researching concentration ratios, it is necessary to fully understand the nature of theexperiment or natural system in which it was measured. The type of plant, the type ofsoil, the quantity of naturally occurring chemical analogues in the soil, and the stage ofgrowth all influence the amount of contaminant absorbed into the plant through the roots.A large fraction of the observed variability may be due to the fact that analytical methodsdo not account for bioavailability of materials in soil - how much of the concentrationmeasured in the soil is actually available to the plant and not bound to the soil matrix. Inaddition, it is important to know whether the concentration measurements are made interms of wet or dry plant. Vegetation is a large percentage water, and the degree ofturgor in the harvested plant has a strong influence in the measured concentration. Theintent in most codes is to use a concentration ratio between dry soil and dry plant matterand to adjust to the full mass of the plant with a dry-to-wet mass ratio. When takingtransfer factors from the literature, it is important to note whether the values are reportedin wet or dry weight. Some researchers differentiate between the concentration ratio forplants consumed as feed by animals (designated as Bv1 and given in units using dryweight of vegetation and soil) and for plants consumed fresh by humans (designated By2and given in units of fresh [or wet] weight of vegetation and dry weight of soil) (Miller etal. 1980). A conversion of these two concepts is given in Peterson (1983) as

B,,2 = By,J ( (1 7)SFW)

where DW is the dry weight and FW is the fresh weight.

Since the handbook prepared by Ng et al. (1968), numerous studies have been undertakento quantify transfer factors (or concentration ratios) for specific chemical elements as afunction of food type. These studies have been compiled in several publications. Mostcomputer codes reference one or more of these compilations as the source of theirtransfer factors. Several frequently referenced compilations include the InternationalAtomic Energy Agency's Technical Report Series #364, Handbook of Parameter Valuesfor the Prediction of Radionuclide Transfer in Temperate Environments (IAEA 1994).This document encompasses a wide variety of plant types and is the result of extensivebackground investigations. It is based on data compiled by the International Union ofRadioecologists. Work is currently underway coordinated by the IAEA to update thisdocument. A second frequently cited reference is the NUREG/CR-5512, ResidualRadioactive Contamination From Decommissioning: Technical Basis for Translating

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Contamination Levels to Annual Total Effective Dose Equivalent (Kennedy and Strenge1992) because of its large set of data and traceable references. Other references includethe National Council on Ionizing Radiation and Protection (NCRP) Report #123 (1996),Screening Models for Releases of Radionuclides to Atmosphere, Surface Water, andGround, and the series of documents by Coughtrey and Thorne (1983), RadionuclideDistribution and Transport in Terrestrial and Aquatic Ecosystems, Vols. 1-6.

The National Council on Radiation Protection and Measurements developed a set of soil-to-plant transfer factors for screening applications (NCRP 1999), basing it largely on theInternational Atomic Energy Agency/International Union of Radioecologists(IAEA/IUR) publication (IUR 1989). However, the NCRP specifically assigneduncertainty ranges to the parameters in the form of geometric standard deviations forassumed lognormal distributions. The assigned geometrical standard deviations (GSDs)ranged from 2.5 to 3.0. However, the transfer factor values adopted by NCRP aredescribed as being "close to the top of the measured range for most nuclides," with highervalues used for sparsely vegetated soils. Because the screening intent was to beconservative, the actual uncertainty range could be larger (with a higher possibility oflower values.) Biwer et al. (2000) adopted the NCRP ranges for use in the RESRADfamily of codes. Beyeler et al. (1999) independently evaluated the CR values for theNUREG/CR-5512 model. They report GSDs ranging from a minimum of 2.47, based on1250 measurements of 23 nuclides as reported by Sheppard and Evenden (1990), to ashigh as 9.5. Most nuclides were assigned the 2.47 default; interestingly, thoseradionuclides with some conflicting data often had higher GSDs than those with no data.Here, the GSD was used to represent experimental variability, not necessarily uncertaintyin an annual average value that is representative of time and space averaging over somedefined area. Somewhat different results were obtained using an expert elicitation process(Brown et al. 1997). It is apparent that there is a large potential variability in the CRvalues as a result of "numerous and complex underlying processes such as climate,growing conditions, plant metabolism, plant rooting traits, soil type, soil moisture, soiltexture, and soil pH (Beyeler et al. 1999)."

Some models make additional assumptions involving root uptake and the surface soillayer when a clean non-contaminated layer of soil is applied over contaminated upper soillayer. The GENII family of codes (Napier et al. 1988; Napier 2003) and RESRADfamily of codes (Yu et al. 2001) also allow the root uptake transfer factor to be modifiedif not all of the plants' roots are in the contaminated soil layer. The GENII and RESRADcodes assume that uptake is directly proportional to the fraction of roots growing in thecontaminated zone. In GENII, this is accomplished using an input root-penetrationfactor, usually set to 1.0 for surface contamination and a lower value for contaminationcovered with a clean surface layer. A similar approach is taken in RESRAD, whichderives cover and depth modifying factors as a function of the thickness of the cleanoverburden above a contaminated layer. Biwer et al. (2000) investigated rooting depthsfor various types of plants, and suggested that for stochastic simulations this be varieduniformly from 0.3 to 4.0 meters. However, Biwer et al. did not consider the actualappropriateness of the assumption of linearity of uptake with root fraction.

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Root uptake is important for biologically active or mobile contaminants; othercontaminants are discriminated against by biological systems, and their root uptake isminimal. For these types of radionuclides, other mechanisms may be the dominant onesfor vegetation contamination. As discussed above, such mechanisms include simplephysical contamination of the external surface of the plant, such as resuspension anddeposition of dust or splashup of contaminated particles of soil during rain or irrigationepisodes. In the recent IAEA recommendations (IAEA 2001), an aggregated transferfactor was developed that implicitly includes the soil adhesion term. To accountimplicitly for soil adhesion, a minimum value of 0.1 is assigned to the transfer factor foranimal feeds and 0.001 to the concentration ratio for humans. The IAEA's (2001)concentration ratio for human foods is given in terms of wet weight. To differentiatethese from the Bv1 and B12 terms defined above, the IAEA gives these the symbols F•1and F•2.

It is important to remember that, for all of the different definitions and applications listedhere, the modeling of transfer of radionuclides from soils to plants is based entirely onempirical observation. For combinations of plant type and soil type that have not beendirectly observed, any factor used is at best an approximation.

The model of Baes (1982) and Sheppard (1985) attempted to correlate soil-to-plantconcentration ratios with the distribution coefficient Kd. The approach, introduced byBaes (1982) and further developed by Sheppard (1985), presents the relationship betweenKd and concentration ratio (CR) in the form

In Kd = a + b (In CR) (18)

where a and b are constants, b having a negative value. Sheppard found that this modelwas generally able to estimate the CR to within about an order of magnitude (perhapsbecause the Kd is an implicit measure of bioavailability). Its use does require ameasurement of the soil Kd, which is simpler to obtain than a site-specific estimate of theCR values; however, it is not specific to various types of crops. A related modelspecifically for cesium, introducing even more unknown parameters related to potassiumconcentrations, was attempted by Absalom et al. (1999). This type of approach haslimited utility and has not caught on in the assessment community. However, it doesindicate that in when quantifying uncertainty using Monte Carlo sampling procedures, thesoil to plant uptake should be correlated with the soil Kd.

3.3.5.2 The Observed RatioA different approach to estimating uptake of radionuclides into plants (and animals, seebelow) is based on the observation that the presence of chemically similar stable elementsin soils can have a significant impact on radionuclide uptake in plants. In soils havinglow concentrations of analogous stable elements, radionuclide uptake by plants may beenhanced, sometimes by orders of magnitude above the levels ordinarily found.Similarly, in soils with high concentrations of analogous elements, uptake may besignificantly depressed. This effect has been characterized by the use of elemental ratiosto predict the transport of radionuclides in the food chain. The ratio of the radionuclide-

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to-stable-element concentration in one medium to the radionuclide-to-stable-element inanother medium (for instance, soils and plants) has been named the observed ratio (OR)(Comar et al. 1957). The OR is defined as

OR = ([Rs] / [Rp]) ([Ap] / [As]) (19)where

[Rs] = radionuclide concentration in sample;[R1] = radionuclide concentration in precursor;[A.] = analogue concentration in sample; and[Ap] = analogue concentration in precursor.

This equation may be rearranged to yield an equation for radionuclide concentration inthe sample based on the observed ratio and the analogue concentrations.

'This approach was developed in the 1950s prior to the concentration ratio approach andwas applied successfully to strontium and its analogue calcium, and somewhat lesssuccessfully to cesium and its analogue potassium. However, it may be made moresuccessful it instead of using concentrations of radionuclides in soil, the concentration insoil solutions is used instead (see discussion below on non-linear responses).

3.3.5.3 Uncommon Food ProductsFarm products commonly included in radiological assessment models include leafyvegetables and produce (NRC Regulatory Guide 1.109 ([NRC 1977]; RESRAD family ofcodes [Yu et al. 2002]), or perhaps leafy vegetables, other vegetables, cereal grains andfruit (used in the GENII family of codes [Napier et al. 1988; Napier et al. 2003] and itsderivatives such as the ERMYN code developed for Yucca Mountain [Wu 2003]). Earlycomputer codes developed from the HERMES model (Fletcher and Dotson 1971) such asFOOD (Baker, Hoenes, and Soldat 1976) used a longer list including

* Leafy above-ground vegetables;* Other above-ground vegetables;• Potatoes;• Other below-ground vegetables;• Berries;* Orchard firuit;* Wheat and wheat products; and* Other grains.

The longer list was reduced because it was observed that soil-to-plant concentration ratioswere not available for most of the crop types, and generic values were being used for allfoods (e.g., Napier et al. 1988).

The limited number of food crops is beneficial to the analyst, because it minimizes thenumber of parameters that must be determined. However, this approach may be seen asethnocentric and biased by members of some groups in the public. In the Columbia RiverComprehensive Impact Assessment (CRCIA) (DOE 1997), which attempted toincorporate public inputs in its development, the lack of specific mention of other foodtypes - and in particular, Native American natural foods - and the lack of a wide rangeof "other" foods was seen as a severe limitation. It was not considered to be useful to

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investigate specific ingestion rates of roots, fruits, etc., unless uptake parameters tospecific plant parts (roots versus leaves) or specific plant species were available.Medicinal and other uses of plant material also provided reasons to increase ingestionrates of specific plant types. The generic approaches to radionuclide uptake in plants areuseful for uncommon foods, but it must be recognized that plant-specific uptake ratios areprobably not going to be available, and uncertainty ranges Mhust be expanded to obtainpublic acceptance.

Following the accidental release of radionuclides at the Chernobyl Power Plant in 1986,the contamination of natural forest products became an issue. Non-cultivated foods suchas berries (noted above) and mushrooms were found to be important. Forests are veryefficient at trapping atmospheric aerosols. The aerosol radionuclides fall on the aerialparts of trees (leaves, branches and trunks), understory, and on the soil (litter, mosses andsoil horizons). Tree canopies can represent large foliar surfaces, up to ten times greaterthan in agricultural areas. The long residence time of radionuclides in forests implies anincrease in potential internal and external doses for decades. The progressivedecomposition of leaves fallen onto the soil constitutes the so-called litter that representsa trap that fixes the radionuclides and slows their migration deep into soil. The largenumbers of mycorhizae and the intense microbial activity in forest soils are the maincauses of cycling of radionuclides and stability of transfer factors for several years afterdeposition. The forest products consumed by humans (mushrooms, berries, so-called"tree juices" such as maple syrup, birch juice) can be up to 250 times more contaminatedthan the equivalent agricultural products 2. The Chemobyl experience has resulted in theavailability of soil-to-plant aggregated transfer coefficients for a few major forestproducts such as berries and mushrooms, for a limited number of radionuclides (primarilynuclides of cesium and strontium). The aggregated transfer coefficient relates the activityconcentration in the forest product (Bq/kg) to the activity of the surface deposit per unitarea (Bq/m2). The unit is then m2 kg"1. Values for aggregated transfer factors can benotably higher than the equivalent ones for agricultural products. It has to be stressed thatit is difficult to get a precise value, because variability is large in such uncontrolledecosystems. Uncertainties of one order of magnitude higher or lower on the estimationsare normal.

The aggregated transfer coefficient for mushrooms is widely variable (up to 3 orders ofmagnitude). This variability is due to several reasons:

* Species plays a role of prime importance. The high transfers of cesium in somemushrooms species could be related to the existence in the carpophore ofsubstances with a great affinity for cesium. Transfer factors have been related tothe species as shown in Table 3.

" The type of mushroom is an indicator of the cesium transfer. Saprophyticmushrooms develop on decomposing materials on the surface of the soil, whichmeans that these kinds of mushrooms will be the first contaminated followingdeposition. Transfer factors will diminish as the deposit migrates into the soil. Thesymbiotic mushrooms associate with trees. Due to their extended mycelium, these

2 p. Calmon, personal communication of data collected for the IAEA EMRAS Program.

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mushrooms bring minerals to the tree and trees provide the mushrooms withorganic substrates from photosynthesis. Most of the edible mushrooms aresymbiotic ones. Unfortunately, this characteristic does not give any informationabout the ability to accumulate cesium by mushrooms. The parasitic mushroomsdevelop at the expense of the trees and are most of the time characterized by lowtransfer factors.Mycelium depth plays a crucial role in the contamination chronology. Mushroomswith surface mycelia will be contaminated quickly after deposition; deepmycelium mushrooms will be contaminated later.

For radionuclides other than cesium, such high transfer coefficients are not observed, noris there as large a variability in the transfer factors - it is generally comparable to that forvegetative plants.

For berries, there is not as large a variability in the transfer factor for cesium as in thecase for mushrooms. However, transfers to berries are higher than those for agriculturalplants.

Soil to plant concentration ratios are being measured for a few radionuclides (primarilystrontium and cesium) for common Asian food types such as rice, Chinese cabbage,radish, and soybean (Choi et al. 1991; 1992; 1995; 1998). Cultivating systems are animportant factor for plant uptake of radionuclides from soil. One of the critical foods forthe intake of radionuclides by humans, especially in the monsoon Asian regions, is rice.Rice is usually planted under flooded conditions because they provide the bestenvironment for the rice plants. No crop rotation is necessary, so that rice has beenplanted in the same places for hundreds of years. Also, the crop can be growncontinuously twice a year in the same paddy fields in tropical or subtropical areas and insome parts of the temperate zones. Because rice is essentially emergent (from the water)vegetation, the main uptake route may not be through the roots. In a tropical area,cropping five times in two years or triple cropping in one year is possible if irrigationwater is available.

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Table 3. Aggregate transfer coefficient for cesium for 4 mushroom classes (Calmon,personal communication of data collected for the IAEA EMRASW Program)

Description Value (m2.kg"l) Mushroom species in the classfresh weight

Mushrooms with Agaricus arvensis, Agaricus sylvaticus, Armillarialow transfer mellea, Boletus appendiculatus, Boletus elegans,coefficient for 1.1072 Cantharellus cornucopiensis, Leccinum aurantiacum,cesium (Macro) Lepiota procera, Lepista nuda, Lepista saeva,

Lycoperdon perlatum, Psalligta campestris, Sarcodonimbr.

Mushrooms with Boletus aestivalis, Boletus edulis, Cantharellusmedium transfer cibarius, Cantharellus palens, Clitocybe nebularis,coefficient for Collybia buthyracia, Collybia confluens, Collybiacesium 5.10-2 dryophylla, Collybia maculata, Collybiaperonata

Hydnum repandum, Kuehneromyces mutabilis,Lactarius deterrimus, Lactarius helvus, Lactariusodoratus, Lactarius picinus, Leccinum sp., Leccinumversipelle, Lepiota naucina, Oudemansiella sp.,Oudemansiella radicata, Pholiota aegerita, Russuladecolorans

Mushrooms with Boletus cavipes, Cantharellus lutescens, Cantharellushigh transfer tubaeformis, Clitocybe infundibuliformiscoefficient for 1.10"1 Lactarius lignyotus, Lactarius quietus, Lactariuscesium torminosus, Lactarius turpis, Leccinum scabrum,

Russula nigricans, Suillus grevillei, Tricholoma aurata,Trichomolopsis rutilans

Mushrooms with Clitocybe cavipes, Dermocybe sp., Hebeloma sp.,very high transfer Hygrophorus sp., Hygrophorus olivaceoalbus,coefficient for Laccaria amethystina, Laccaria laccata, Laccariacesium 5.10"1 proxima, Lactarius sp., Lactarius camphorates,

Lactarius necator, Lactarius porninsis, Lactarius rufus,Lactarius theiogalus, Lactarius trivialis, Rozitescaperata, Russula sp., Russula badia, Russulaerythropoda, Russula ochroleuca, Russula turci, Suillusbovinus, Suillus granulatus, Suillus luteus, Suillusvariegatus, Xerocomus badius, Xerocomuschrysenteron, Xerocomus subtomentosus

(Data from Amundsen, Gulden, and Strand 1996; Baeza et al. 2000; Bakken and Olsenl990;Barnett et al. 1999; Battiston et al. 1989. Borio et al. 1991; Byrne 1988; Djingova and Kuleff2002; Fraiture, Guillitte, and Lambinon 1990; Gaso et al. 1998; Gaso et al. 2000; Giovani, Nimis,and Padovani 1990; Grueter 1971: Guillitte, Fraiture, and Lambinon 1990; Heinrich 1992;Heinrich 1993; Horyna and Randa 1988; Kammerer, Hiersche, and Wirth 1994; Kirchner andDaillant 1998; Lamnbinon et al. 1988; Mascanzoni 1990; Mietelski et al. 2002; Pietrzak-Flis et at.1996; Randa et al. 1990; Rantavaara 1990; Roemmelt, Hiersche, and Schaller 1988; R6mmelt etal. 1990; Riihm et al. 1998; Seeger and Schweinshaut 1981; Steiner, Linkov, and Yoshida 2002;Svadlenkova, Konecny, and Smutny 1996; Tsukada, Shibata, and Sugiyama 1998; Vinichuk,Johanson, and Taylor 2003)

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3.3.5.4 Potential Non-Linear UptakesA key assumption of the concentration-ratio approach is that the CR is a constant as afunction of concentration; that is, the uptake is linear. This may not be true forcontaminants that are nutrients or are chemical analogues for them. Non-linear responsesmay be possible if plants scavenge essential elements at low concentrations but maintaina homeostatic balance at higher soil concentrations (Sheppard and Sheppard 1985). Theassumption of linearity may be appropriate for elements that are not essential tobiological function, are not analogues of such elements, or are not absorbed by organismsvia nutrient pathways. These latter elements seldom exhibit linearity at very lowenvironmental concentrations (Figure 3). For such elements (Table 4),. organisms oftenare able to homeostatically regulate their tissue concentrations over a range ofenvironmental concentrations. In some cases, these concentrations may be 40 to 200times greater than the amount needed to sustain life (F6rstner and Wittmann 1981).There is generally little information to evaluate this concept except for a fewradionuclides such as cesium, which mimics potassium, or strontium, which mimicscalcium.

The differences in predicted effects between these two models can be substantial.Experience in assessing risks of metals in sediments and groundwater indicates that, forcertain metals and species, the linear model can overestimate exposure by up to six ordersof magnitude (DOE 1997).

I

01

Figure 3. Relationships are shown between tissue concentrations and environmentalconcentrations for non-nutrient and nutrient compounds (adapted frompersonal communication with C. A. Brandt).

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Table 4. In a biological classification of metals (modified from Beeby 1991) shaded cellsare non-nutrient analogs; bold cells are nonessential metals.

Period J Macronutrient J Micronutrient Non-essential

3 Na Mg

4 K Ca Cr Mn Fe, CU ZnCo, Ni

5 '

6 Hg Pb

7 Eu, U Np

(a) Period is from the Periodic Table of the Elements.

To date, few models have been developed to deal with this problem. A possiblealternative for radionuclides with this behavior would be to treat them with a specific-activity type model, wherein the atom ratio of the radioactive to stable atoms of anelement are assumed to be the same in the plant as in the soil (see below). The modelproposed by Norden et al. (2005) is essentially a specific-activity type model; theysuggest that transfer rates of contaminants would mimic analogue nutrient transfer rates.Their suggestion is that for the case of radionuclides that are analogues of macronutrients,it can be assumed that the plant uptake of the nutrient modulates the uptake of theradionuclide. This means, that the radionuclide and the corresponding analogue nutrientare taken up by plants in an identical manner. Furthermore, assuming that only ions in thesoil solution near the roots are available for transition into the roots, the transition of theradionuclide from soil to plant roots can be represented as an independent Poissonprocess with the following rate:

TR(RN) = RN] TR(A) (20)

[RN] + [A R(

where,TR(RN) is the transition rate of the radionuclide from the soil solution into the roots

[mol/yr],TR(A) is the overall transition rate of the analogue from the soil solution into the

roots [mol/yr][RNj],, is the radionuclide concentration in the soil solution near the roots [mol/m 3][A],, is the analogue nutrient concentration in the soil solution near the roots

[MolI/m 3]].Because [RN]. is generally very small with respect to [A].s, the overall transition ratefrom the soil solution into the roots equals the analogue uptake rate. These authors alsosuggest that a selectivity coefficient may be added to this equation if the radionuclide hasa slightly lower uptake than the analogue. This model is very similar in concept andapplication to the Observed Ratio, except that it is based on soil solutions rather than soilconcentrations.

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3.3.5.5 Specific-Activity ModelsSome key biologically-active elements are modeled in a sepaiate way through the use ofspecific activity models. The behavior of the radionuclides tritium and 14C in exposurepathways must be handled in a special manner, because the behavior of theseenvironmentally mobile radionuclides mimics that of stable hydrogen and carbon inbiological processes, respectively. The concentrations of tritium or '4C in environmentalmedia (soil, plants, and animal products) are assumed to have the same specific activity(curies of radionuclide per kilogram of soluble element) as the contaminating medium (airor water). The fractional content of hydrogen or carbon in a plant is then used to computethe concentration of tritium or 14C in the food product under consideration. The hydrogencontent in both the water and the nonwater (dry) portion of the food product is used whencalculating the tritium concentration. For airborne releases, it is assumed that plants obtainall their carbon from.airborne carbon dioxide.

In the simplest implementation, tritium in the environment is assumed to be in the oxide(HTO) form, and to behave in a manner identical with water. For HTO releases,concentrations in air moisture (Bq L-1) are found by dividing the concentration in air (Bqm73) by the absolute humidity (kg or L m-3). The concentration of tritium in vegetation iscalculated (Baker, Hoenes, and Soldat 1976) as

CTp = 9 Cw FHp (21)

where CTp = the concentration of 3H in plant type p (Bq/kg);CwT = the concentration of tritium in the environmental water (Bq/L); andFHp = the fraction of hydrogen in total vegetation of plant type p.

The coefficient, 9, converts tritium concentration in environmental water to concentrationin hydrogen (this is approximate, but applicable to water with dilute solutions of HTO).Note that I kg of water equals 1 L in the above equation. The fractions of hydrogen invarious food types are given in Table 5. (Some additional parameters are included in Table4 for the discussion of animal products given below).

Murphy (1986) and Raskob (1994) have developed models designed specifically tocalculate doses from chronic releases of elemental tritium (HT) and HTO, taking intoaccount the formation of organically-bound tritium (OBT). Both models are similar inthat they model dispersion of HT and HTO, deposition of HT and HTO through drydeposition, and deposition of HTO through wet deposition. Treatment of the soilcompartment includes such processes as diffusion and resuspension. In NORMTRI(Raskob 1994), air concentrations due to resuspension are calculated by treating the areaof deposition as an array of ground-level area sources. For an HT release, 20% of thedeposited tritium is assumed to remain in the soil; the rest is re-emitted. In the Murphymodel, HT is oxidized in the leaf as well as in the soil. Both models account explicitlyfor the transfer of OBT in plants to OBT in animal products. However, both of thesemodels require extensive information and operate on a limited time scale. Forassessments of a more general nature, the model of Peterson and Davis (2002) thatincludes both HTO and HT is useful.

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Table 5. Weight Fractions of Hydrogen and Carbon in Environmental Media, Vegetation,and Animal Products (Napier et al. 2003)

Food or Carbon Hydrogen Carbona Hydrogenb

Fodder Water dry dry wet wet

f. f, fh F.,, F. Fhv, Fha

Fresh fruits, 0.80 0.45 0.062 0.090 0.10vegetables,and grassGrain and 0.12 0.45 0.062 0.40 0.068stored animalfeedEggs 0.75 0.60 0.092 0.15 0.11Milk 0.88 0.58 0.083 0.070 0.11Beef 0.60 0.60 0.094 0.24 0.10Poultry 0.70 0.67 0.087 0.20 0.10a F,, or Fca = fc (1 - fL). F applies to plants and F, applies to animal products.b Fhu or Fha = (f. /9) + fh (I - fC). Fh, applies to plants and Fha applies to animal products.

To estimate HTO concentrations in plant products, Peterson and Davis (2002) assumedconservatively that the concentration of HTO in plant water equals 0.9 times the HTOconcentration in air moisture (Cm) for leafy vegetables and pasture and 0.8 Cam for fruit,root crops and other vegetables, and grain. These reduction factors account for dilutionfrom soil water, which has a lower tritium concentration than does air moisture andwhich affects fruit and root crops more than leaves. OBT concentrations are estimatedfrom a plant water concentration of 0.9 C. for all plant types, since OBT is assumed toform exclusively in the leaves and to be translocated to other plant parts. Then, isotopicdiscrimination occurring in the formation of OBT (Table 6) is assumed conservatively toresult in an OBT concentration 0.9 times the concentration of HTO in plant water.

The Peterson and Davis (2002) equation for calculating concentrations of HTO in plantproducts (Bq kg-1 fw) is:

CppHTo = RFpp Cam Ff_.pp (22)

whereCppýHTO = concentration of HTO in plant products (Bq kg"1 fresh weight)Ff,,__pp = average fresh weight (fw) fraction of plant products (kg kg") (Note, this

is essentially the same as fw in the preceding equation)RFpp = reduction factor that accounts for low soil water concentrations relative

to concentrations in air moisture and varies by vegetable typeCam = HTO concentration in air moisture (Bq L-')

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Table 6. Parameter values for calculating HTO and OBT concentrations in plant products(adapted from Peterson and Davis 2002).

Leafy Fruit Root and Grain Pasturevegetables other

vegetablesReduction factor (RFpp) 0.9 0.8 0.8 0.8 0.9Reduction factor for leaves (RF,) 0.9 0.9 0.9 0.9 0.9Isotopic discrimination factor (IDpp) 0.9 0.9 0.9 0.9 0.9Fresh matter fraction (Ffw,_pp) 0.906 0.853 0.824 0.117 0.8Dry matter fraction (F&_.,p) 0.094 0.147 0.176 0.883 0.2Water equivalent factor (W i,,,) 0.6 0.59 0.58 0.577 0.616

The Peterson and Davis (2002) equation for calculating concentrations of OBT in plantproducts (Bq kg' fw) is:

Cpp_oBT= RF, IDpp Cam Fdm._p Weqpp (23)where

RFi = reduction factor for plant leavesIDpp = isotopic discrimination factor for plant productsCpp OBT = concentration of OBT in plant products (Bq kg"' fresh weight)Fdm_p~p = average dry matter fraction of the plant products (kg dry mass kg'1 fresh

mass)WqLpp = water equivalent factor of plant products (L kg')

HT is a biologically inert gas and imparts an extremely low dose relative to tritiatedwater. However, HT is converted in the environment to HTO, in which form it movesthrough the food chain to dose. Therefore, a model of dose from HT releases isnecessary. Ratios of the HTO concentrations in soil, air and plants to the HTconcentration in air recommended for screening level assessments by Peterson and Davis(2002) are 8 Bq L" / (Bq m"3) for the ratio of HTO in air moisture to HT in air, and 12 BqL" / (Bq nf) for the ratio of HTO in plant water to HT in air. Doses from HT are thenmodeled using the recommended HTO / HT ratios for air moisture and plant water.Concentrations in air moisture are calculated by multiplying the HT concentration in air(Camf; Bq 1"3) times eight. Concentrations of HTO and OBT in plant and animalproducts are calculated for releases of HT by replacing RFpp in the above equations witha factor of 1.5. Thus the plant water concentration in these equations becomes 1.5 C. =12 Ca.

Using the specific activity model for 14C in air, the concentration of 14C in crops fromatmospheric contamination Cc, is calculated (Baker, Hoenes, and Soldat 1976) as

Ccp= Cgc Fcp/Pc (24)

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where Cac = the concentration of 14C in air (Bq/m3);Fcp = the fraction of carbon in the plant (dimensionless); andPc = the concentration of carbon in air (kg/m3).

Plants obtain most of their carbon via photosynthesis of CO2 from the air. Therefore, it isdifficult to estimate the transfer of radioactive 14C from irrigation water or soil to plants.The approach adopted by Napier et al. (1988) is to assume that plants obtain a smallfraction of carbon from soil, and to use that activity ratio. The approach used by Yu et al.(2001) is to assume that carbon in the soil slowly converts to CO 2 and to then estimate a'4CO2 concentration in the air near crops. Due to the volatility of 14C in soil, it is quicklyreleased via gaseous emission as 14CO 2. Sheppard et al. (1991) measured the rate of 14C

loss from soil in outdoor lysimeter experiments. Carbon loss from the soil, measured bythe emission rate (the "evasion rate" in RESRAD), is 12/yr for clay and loamy soils, and22/yr for sandy and organic soils (Yu et al. 2001). Thus, 14C concentrations in surfacesoil reach equilibrium within 1 to 2 months. Emission is the dominant mechanism forremoving 14C from the soil. In comparison, losses due to leaching, radioactive decay, andsoil erosion are slight.

Specific activity models have been recommended for other radionuclides. Attributes thatlead to this type of recommendation are long half-life, high mobility, biologicallyessential element, and massive isotopic dilution in the geosphere and biosphere. "291 and36CI are some of the most critical radionuclides to consider in intermediate- and high-level nuclear-waste management, and both have these attributes. Uptake in plants can besimulated using equations similar to those above for tritium or 14C, using the measuredconcentrations of the stable elements in local soil and plants. The specific-activityapproach has also been suggested for 35S; however, its half-life is only about 3 months,and so is unlikely to be present in radioactive wastes.

3.3.5.6 Uptake in Fruit TreesThe standard uptake model described above using concentration ratios was originallydeveloped for annual crops (leafy vegetables, root vegetables, forages, and grains). Thevast majority of available observations, either laboratory or field studies, are for a limitednumber of crop types. In addition, observations are not available for all crop types for allchemical elements, and most compendiums of transfer factors use surrogates, relying onsimilar plant types, soil types, or chemical behavior to fill gaps in the knowledge (e.g.,Staven et al. 2003; Yu et al. 2001).

The standard model is also used for perennial plants, including fruit trees. The modelmay not be appropriate for trees, considering their longer life and potential foraccumulating contaminants in roots, trunks, and leaves, with transport to fruits possiblydelayed for periods of over 1 year. The British Ministry of Agriculture, Fisheries, andFood (MAFF, since 2001 called the Department for Environment, Food, and RuralAffairs) has noted that such "models for fruit are extremely conservative... Extensiveresearch is underway to produce more appropriate data. However, a review for MAFFconcludes that no better fruit models currently exist" (MAFF 1999).

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The International Union of Radioecologists and the International Atomic Energy Agencyjointly prepared a major review of soil-to-plant transfer factors (IAEA 1994), which isused as a basic reference world-wide. In this review, soil-to-plant transfer factors arepresented for many radionuclides and crop types. None of them refer to fruit or nut trees.A more recent compilation for tropical ecosystems does include some fruit trees(primarily apples), but the only nut is the coconut (IAEA-IUR 1997, quoted in Carini2001). This state of affairs has been discussed in the international arena for several years;the IAEA's Validation of Model Predictions (VAMP) Program Multiple PathwaysAssessment Working Group noted in the early 1990s that many participantsoverestimated concentrations of 137Cs in fruit trees, and that models for predicting thecontamination of fruit were in need of further improvement (IAEA 1995). The sequel toVAMP, the IAEA's Biosphere Modeling and Assessment (BIOMASS) Program,included a Fruits Working Group (IAEA 1996). As part of the work of the FruitsWorking Group, a review was undertaken of the experimental, field and modelinginformation on the transfer of radionuclides to fruit. The results of this work werepublished as a special issue of the Journal of Environmental Radiation (Ventner 2001).

In this special issue, Mitchell (2001) reports that there are three generic types of modelsthat are applicable to fruit trees:

* Simple mathematical functions describing declining concentration in fruit basedon observations following deposition (e.g., Antonopoulos-Domis et al. 1990);

" Models that attempt to predict temporal distribution in soil-plant systems throughdescriptions of the processes involved, e.g. a model postulated by Frissell at the1994 VAMP meeting in Vienna;

" Radiological dose assessment models that use a mixture of equilibrium and/ordynamic modeling approaches to predict concentrations in edible products, e.g.SPADE (Thorne and Coughtrey 1983).

Antonopoulos-Domis et al. (1990) developed a model structure forperennial fruit treesdescribing distribution, retention, transfer and rejection of radionuclides, based onexperimental determinations of 13 7Cs in apricot fruit trees. The original concept for themodel was based on the fact that the leaves and fruits developing each year are onlycontaminated by a portion of the 137Cs in the body of the tree. A fraction of this availablereservoir is removed each year, part is lost from the tree through leaves and fruit and partbecomes irretrievably associated with the body of the tree. This model requiresknowledge of the contaminant inventory in the soil and the tree, as well as the deposition.The model is not immediately transferable to other types of trees, radionuclides, andlocations, but it does indicate that at least two compartments are probably necessary toadequately describe the long-term accumulation and transfer of contaminants from treesinto fruits.

A model for radiocesium transfer to tree fruit described by Frissel (1994) considers thehomeostatic control of potassium within fruit trees. The model structure has fourcompartments and was designed to consider the long-term fate of cesium in soil asaffected by changes in the supply of potassium to soil. The four compartments are soil,the easily accessible part of a tree, the poorly accessible woody part, and the fruit or leaf.

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The model is homeostatic, i.e. all cesium concentrations and fluxes are controlled bypotassium concentrations and fluxes, respectively. In determining the various transferparameters, it is assumed that there is no difference in the behavior of potassium andcesium, but that discrimination occurs between the compartments. The loss of plantmaterial, termed debris by Frissel, via branches, leaf fall and fruit loss is included andreturned to the soil, but because uptake is homeostatically controlled, this has minimalinfluence on the tree contents. Frissel (1994) concluded that the model was probably notsufficient to describe cesium transfer to fruit. In particular, the use of three compartmentswas not sufficient to model availability within the plant. The model results do indicatethat important processes are likely to be the biological half time of cesium in wood, thediscrimination between cesium and potassium, cycling of potassium (through fallingleaves, etc.), and uptake of potassium.

The fruit plant model in the SPADE computer code (Thorne & Coughtrey 1983) has sixcompartments, representing internal leaf, external leaf, stem, fruit, storage organs androot. Movement of radionuclides within the plant model is controlled by empiricallyderived rate constants and parameters are derived for three broad categories of fruit plant:herbaceous, shrub and tree. Foliar absorption is represented by transfers between theexternal leaf and internal leaf compartments. Interception by plants takes account ofchanges in plant biomass with season. The original default parameters were basedlargely on data for cereals but were modifed in the case of tree and shrub fruits to allowfor more rapid transfer from stem to root so that the root store could serve as a reservoirthrough subsequent seasons. Loss of radionuclides from external plant surfaces to the soilis modeled as transfer to the surface layer of the soil model. The process of root uptake ismodeled as the transfer of radionuclides from soil solution to the plant root compartment.The transfer rate is also assumed to vary with soil layer depth, both as a function of theroot distribution throughout the soil profile and as a function of the deposit distribution insoil. Consequently, the transfer of radionuclides from the soil solution to root isrepresented by a discrete transfer from each of the 10 layers in the soil model. The soilsolution to root rate constant in each soil layer is a function of the root uptake rateconstant and the assumed distribution of root activity in each layer. Three plantabsorption mechanisms are responsible for the transfer of radionuclides at the soil-rootinterface: plant-base absorption, main root system absorption and tap root absorption.The actual value of the soil solution to root transfer coefficient for each root layercorresponding to the soil layer of the soil model is calculated as the product of thespecified rate coefficient and the normalized root shape modifier.

None of these models appears to be suitable for generic use in long-term radiologicalassessments without substantial modification and simplification, and all requireadditional development of parameters before general use. Again, it was noted thatmeasurements of radionuclide uptake in trees were lacking (Carini 2001), andrecommended that "There is a need for research on the behavior of radionuclides in fruitcrops to drive model development, not simply to parameterize existing models. Researchshould focus on understanding the key processes" (Coughtrey et al. 2001). As a result ofthese recommendations, the IAEA initiated the Environmental Modeling for RadiationSafety (EMRAS) program in 2003. This program has a Working Group on Revision of

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IAEA Technical Report Series No. 364 "Handbook of parameter values for the predictionof radionuclide transfer in temperate environments" and attempted to establish anotherFruits Working Group. However, the participation in the Fruits group was so low thatthese participants joined with the Handbook group.

3.4 Modeling the Uptake of Radionuclides in Animal ProductsIn general, animals and animal products may become contaminated through a widevariety of pathways, in ways similar to humans. Pathways by which animal productsmay become contaminated include animal ingestion of plants, water, and soil (theanimals' inhalation is usually neglected as contributing little to the total as discussedbelow). However, as with people, the source of the foods and water may be quite varied,and include contributions from contaminated soil as well as, for predators, otherpotentially-contaminated animals.

Absorption of radionuclides by animals via inhalation is of limited importance. Hvindenet al. (1964) showed clearly that the intake of radionuclides through inhalation was atleast three orders of magnitude less than that from pasture. Since the respiratory tract isnot usually an edible product for man, the significance of the inhalation route for foodchain contamination is further reduced.

3.4.1 Feed to Animal Product Transfer FactorsThe various routes of ingestion by animals (plants, water, prey) can be combined throughthe use of a transfer coefficient. The transfer coefficient is widely used in the literature topredict the transfer of radionuclides to animal food products. Ward and colleagues firstused the transfer coefficient to describe the relationship between 137Cs in animal feed andits concentration in milk of dairy cows (Ward and Johnson 1965; Stewart et al. 1965;Johnson et al. 1968). They subsequently described the validation of this term asconceived (Ward and Johnson 1986). Ward and colleagues expressed their results as aratio of the daily 13Cs intake (Bq/day) to the concentration of Cs per liter of milk(Bq/L). This term was used to successfully describe the relationship of 137Cs levels infeed and milk under varying environmental management conditions. Subsequently, thisapProach was found to provide a convenient method to describe the relationship between1"Cs in feed and meat products (Johnson et al. 1969).

A detailed mathematical derivation of the animal product transfer function is provided inPeterson (1983). Much of the early development of this function was done by Ng andassociates (Ng and Thompson 1968; Ng et al. 1977; Ng, Colsher, and Thompson 1979;Ng 1982; Ng, Colsher, and Thompson 1982a; Ng, Colsher, and Thompson 1982b).This relationship (Fm) has subsequently been expressed in the literature by the followingequation:

Radionuclide Milk Concentration (Bq / L)Feed Concentration (Bq / kg) x Feed Intake (kg / day)

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For appropriate estimations, the feed concentration and the feed intake should be long-term averages of what is eaten rather than point measurements. Since the inception ofthis concept, transfer coefficients have been applied to many of the radionuclides;nevertheless, Ward and Johnston (1986) cautioned regarding the general applicability ofthe use of transfer coefficients for a wide number of radioelements and animal species.Others argue about the use and abuse of transfer coefficients, but according to Shaw(2001), "no one has yet come up with a better idea" for expressing this relationship. Thesame relationship is used for other animal products such as meat and eggs.

Transfer to animal products has been expressed using various quantitative terms,sometimes with different dimensional units. These have included terms such asbioconcentration factor, transfer coefficients, transfer factor, concentration factor,concentration ratio, and aggregated transfer coefficient (ICRU 2001). In an effort tosimplify and standardize this widely differing terminology, the International Commissionon Radiation Units and Measurements has adopted a broad term called the concentrationratio, or Cr, which is defined as the ratio of the activity density of a radionuclide in thereceptor compartment to that in the donor compartment (ICRU 2001). In the case of plantto animal transfers, the concentration ratio in this case is defined as the steady-stateconcentration in an animal tissue divided by the average concentration in feed.

The concept of the transfer factor may be applied to essentially all routes of ingestion byanimals: soil-to-animal, plant-to-animal, water-to-animal, and prey animal-to-predatoranimal.

Some dynamic models such as PATHWAY( Whicker and Kirchner 1987) useassimilation fractions for ingested materials and biological elimination rates; however,these are also directly derived from the steady state or equilibrium transfer factors.

As with radionuclide uptake into plants, some elements in animals may behomeostatically controlled. Very little work has been done to quantify the effect that thismight have on animal product transfer factors.

3.4.2 The Observed Ratio in Animal Products

A different approach to estimating uptake of radionuclides into animals is based on theobservation that the presence of chemically similar stable elements in feed can have asignificant impact on radionuclide uptake in animals. The ratio of the radionuclide-to-stable-element concentration in one medium to the radionuclide-to-stable-element inanother medium (for instance, feed and animals) has been named the observed ratio(OR). Early work developed the OR to predict the transfer of 9°Sr uptake by milk (Comaret al. 1961; Comar et al. 1966). In this sense, the OR is described as the ratio ofstrontium to calcium in the milk following consumption of 9°Sr-contaminated feed.Likewise, the cesium/potassium ratio was used to describe the uptake of 137Cs in milk,assuming that the relationship would be similar for cesium and potassium as forstrontium and calcium.

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Although this approach has not been widely used for many years, it could potentially beused to simulate the uptake of radionuclides that are analogues of stable elements that arehomeostatically regulated in animals. Homeostatic control via major routes allows theanimal to regulate the tissue content of the various elements taken into the body, mostlyvia ingestion. These control systems include gastrointestinal absorption, urinaryexcretion, tissue deposition (storage pools or organ uptake such as the thyroid), secretioninto milk, and endogenous excretion via feces. Some radionuclides have dietary essentialstable isotopes (such as iodine) or analogues (such as calcium in the case of strontium),which are under homeostatic control by the animal. For other nuclides, control may beby the mechanisms controlling other essential elements. The intestine, thyroid gland, andmammary gland all have unique mechanisms for controlling many elements. Forelements that are homeostatically controlled, the uptake and excretion is set to maintainthe plasma or milk concentration. In this case, the transfer coefficient will not beconstant over a wide range of conditions, and milk or tissue concentration may not varyas the intake changes. An increase in feed concentration of a radionuclide may notequate to an increase in tissue concentration for these contaminants.

3.4.3 Unusual Animal Products

The animal products commonly modeled include meat, milk, poultry, and eggs (e.g.,GENII [Napier et al 2003], ERMYN [Wu 2003]). Beef is frequently the representativeanimal product for all meat (including beef, pork, wild game, and other meat), cows arethe representative milk-producing animal, and chicken is the representative poultryanimal and egg producer. Some compilations of transfer factor values are available forother animals, such as pork and lamb (e.g., Ng, Colsher, and Thompson 1982). As ageneral rule, transfer factors are not explicitly available for other meat animals such asdeer, goats, rabbits, wild game birds, or other animals occasionally consumed by humans.A limited amount of data are available for transfer of contaminants to goat's milk(primarily for iodine). In an appendix to their report, Ng, Colsher, and Thompson (1982)provide information on scaling reported values of the transfer factor for animals withcharacteristics different from those of the reference livestock. Such information could beused to scale estimates of transfer factors for other types of animal -product, such as eggsfrom birds like ostriches or emus (see discussion below on allometric approaches).

Arctic mammals such as seals, walrus, and whales are a unique subset of animalsconsumed by humans. A review by Layton et al. (1997) provides transfer factors for afew natural and artificial radionuclides for these types of animals.

Some nomadic societies consume products from herded sheep and horses. Some limiteddata are available for actinides and long-lived fission products in these animals frommeasurements made at the Semipalatinsk Test Site in Kazakhstan. The SemipalatinskTest Site, known as the "polygon", is a 19,000 km2 zone located in the northeast ofKazakhstan, 800 km north of the capital Almaty. Between 1949 and 1989, the formerUSSR conducted about 456 nuclear explosions at STS. Until 1963, most of theexplosions were carried out on the surface and in the atmosphere (including 126atmospheric tests with 30 surface bursts). During the 40 years of testing, the total energyreleased in the testing was equivalent to 17.4 megatons of TNT. After the breakup of the

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Soviet Union, effective control over the area by local governments was lost. The regionis a combination of arid plains and lightly forested mountains. Between 30,000 and40,000 people live near the site, and a small number of people actually live on the site,primarily semi-nomadic herders and farmers. Following the breakup of the Soviet Union,the Russian military moved out of the STS, and took nearly all records of activities andresulting contamination with them. The National Nuclear Center, Institute of RadiationSafety and Ecology (IRSE), in the city of Kurchatov, Kazakhstan, carries out research onradiation safety at the STS and radioecology of plants and animals. Preliminary contacts(personal communication, Dr. Larisa Ptitskaya to Bruce Napier) indicate that informationmay be available for the radionuclides of tritium, cobalt, strontium, cesium, europium,plutonium, americium, and radon.

A few other animal products consumed by humans are not amenable to estimation usingtransfer factors. An example is honey from honey bees (e.g., Bonazzola et al. 1991;Fulkery et al. 1998). Accumulation of radionuclides in honey has been reported;however, no models for estimating concentration from environmental contaminationlevels are known. Bonazzola et al. (1991) reported that honey can be contaminated byradionuclides accumulating in soils or plants. The primary pathway of contamination isplant nectar; pollen did not appear to be a dominant contributor. However, theradionuclide content did not appear to be correlated between radionuclides. In addition,the plant concentration varied sufficiently that multifloral honeys could not be traced.These researches.concluded that consideration of the botanical sources of the honey wasalso insufficient to allow use of honey as a bioindicator.

3.4.4 Specific Activity Models for Animal Products

Some key biologically-active elements are modeled in a separate way through the use ofspecific activity models. The behavior of the radionuclides tritium and 1

4C in exposurepathways must be handled in a special manner, because the behavior of theseenvironmentally mobile radionuclides mimics that of stable hydrogen and carbon inbiological processes. The concentrations of tritium or 14C in environmental media (soil,plants, and animal products) are assumed to have the same specific activity (curies ofradionuclide per kilogram of soluble element) as the contaminating medium (air or water).The fractional content of hydro&en or carbon in an animal product is then used to computethe concentration of tritium or 1 C in the food product under consideration. The hydrogencontent in both the water and the nonwater (dry) portion of the food product is used whencalculating the tritium concentration. For airborne releases, it is assumed animals obtain alltheir carbon through ingestion of plants.

The simplest model for tritium in animal products (Baker, Hoenes, and Soldat 1976)defines the concentration of tritium in the animal product as

CTP Q, + CT, Q,(CT,, -F (26)Flv Qf + Q.,/ 9

where CT, = the concentration of 3H in animal product (Bq/kg or Bq/L);CTp = the concentration of 3H in animal product feed crop p (Bq/kg);

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CTw = the concentration of 3H in animal drinking water (Bq/L);Fhf = the fraction of hydrogen in animal feed (dimensionless);Fha = the fraction of hydrogen in animal product m (dimensionless);

Qf = the quantity of feed consumed per day (kg/d); andQw = the quantity of water consumed by the animal per day (Lid)

9 = coefficient converting concentration of tritium in water to concentrationof tritium in hydrogen, valid for dilute solutions.

For the variant of this model proposed by Peterson and Davis (2002), the generalizedequations for calculating HTO and OBT concentrations in animal products (Bq kg"' fw)are:

Ca, Irro = [RFpp Fft + RF1 IDpp Dfr + Wfr RFd. + ISAfr] C. Ffr q, (27)

whereFfr = water fraction contributed to the diet by the water fraction of food;Dfr = water fraction contributed to the diet by the dry matter fraction of food;Wfr = water fraction contributed to the diet by drinking water;RFdw = fraction of HTO concentration in drinking water relative to that in air

moisture;ISAf& = water fraction contributed to the diet by inhalation and skin absorption,

and the other parameters are as defined for plants.

The equation to calculate OBT in animal products assumes the specific activity in organicmaterial equals the specific activity in the aqueous phase apart from a discriminationfactor (Weqa.p).

CC..T = [RF;, Fr + RF, IDpp Dfr + Wfr RFd + ISAfI C. F,_ W,_, (28)

Recommended default values for parameters in these equations are presented in Table 7.Values for fresh and dry matter fractions and for water equivalent factors are from Geigy(1981).

As with Peterson and Davis' (2002) model for HT in plants, the ratios of the HTOconcentrations in soil, air and plants to the HT concentration in air recommended forscreening level assessments by Peterson and Davis (2002) are 8 Bq L" / (Bq nm" ) for theratio of HTO in air nioisture to HT in air, and 12 Bq L 1 / (Bq m" 3) for the ratio of HTO inplant water to HT in air. Doses from HT are then modeled using the recommended HTO/ HT ratios for air moisture and plant water. Concentrations in air moisture are calculatedby multiplying the HT concentration in air (Can; Bq m"3) times eight. Concentrations ofHTO and OBT in animal products are calculated for releases of HT by replacing RFvp inthe above equations with a factor of 1.5.

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Table 7. Suggested parameter values for calculating HTO and OBTanimal products (adapted from Peterson and Davis 2002).

concentrations in

Eggs Milk Beef Pork PoultryFraction water ingested from fresh 0.034 0.371 0.409 0.031 0.034weight food (Ffr)Fraction water ingested from dry 0.149 0.065 0.074 0.135 0.149weight food(Dfr)Fraction water ingested from 0.781 0.544 0.490 0.782 0.781drinking water (Wfr)Fraction water from inhalation and 0.036 0.021 0.028 0.052 0.036skin absorption (ISAfr)Drinking water HTO / air moisture 0.5 0.5 0.5 0.5 0.5HTO (RFdW)

Fresh weight fraction (Ffwap) 0.74 0.897 0.668 0.50 0.67Dry matter fraction (Fdcmap) 0.26 0.103 0.332 0.50 0.33Water equivalent factor (W, ap) 0.835 0.669 0.795 0.904 0.796

The concentration of 14C in animal products is calculated as

Co, Qf + C". Q.,CcaFF F,.

.f ~Qf+ Fý.Q.(29)

where C.CCPCcwFcfFcw

QfQw

= the concentration of 14C in animal product (Bq/kg or Bq/L);= the concentration of 14C in crop used for animal feed (Bq/kg);= the concentration of 14C in animal drinking water (Bq/L);= the fraction of carbon in animal feed (dimensionless) (Fc, in Table 5);= the fraction of carbon in animal drinking water (dimensionless);= the fraction of carbon in animal product m (dimensionless) (Table 5);= the quantity of feed consumed per day (kg/d); and= the quantity of water consumed by the animal per day (L/d).

3.4.5 Kinetic Modeling for Animals

Animals may be exposed to a contaminant via ingestion of contaminated water, soil, andfood. The ECEM (Ecological Contaminant Exposure Model) is a food-web basedecological exposure assessment tool (Eslinger et al. 2004). ECEM is a detailed, food-webbased, chronic exposure model intended for use in situations where an organism'sexposure to contaminants in the environment are of sufficient duration to result incontaminant uptake by the organism. This model was developed to allow estimation ofconcentrations and doses to biota for both hazardous chemicals and radionuclides;however, its structure allows its use in estimating concentrations in animals that humansmight consume as well. The ECEM equations to estimate tissue concentration aredescribed below for ingestion (similar sets of equations are also used to estimate animal

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exposures from dermal contact and inhalation - these are not regularly used forradionuclides). The basic ECEM structure is predicated on equilibrium absorption andloss rates from the animal.

Water Ingestion: The equilibrium contaminant tissue concentration in species i fromingestion of water is calculated from the following expression:

C.,= D. (a;ie (30)

whereCgwi = tissue concentration in species i from ingestion of water (pg/kg wet tissue

or pCi/kg wet tissue)Dijn•i = applied daily dose to species i from water ingestion (pg/kg wet tissue/day

or pCi/kg wet tissue/day)aiwatr= ingestion absorption factor for species i from water ingestion (unitless)K, = loss rate of contaminant for species i, including depuration and metabolism

(i/day).

The applied daily dose to species i from water ingestion is calculated from the equation(modified from EPA [1993]):

D. =[VEC., W9,mwji I (31)

whereDing.i = applied daily dose to species i from water ingestion (pg/kg wet tissue/day

or pCi/kg wet tissue/day)ECwýatr = concentration in surface water (pg/L or pCi/L)WIj = water ingestion rate of species i (L/day)Wi = body weight of species i (kg wet tissue)0, = area use factor for species i (ratio of contaminant area to home range -

unitless)Wi = seasonality factor for species i (fraction of year spent at the contaminated

site - unitless)

Soil Ingestion: The contaminant body burden to species i from ingestion of soil iscalculated from the equation:

Cw ( J, ) (32)

where

Cingsi = contaminant tissue concentration in species i from ingestion of soil (pg/kgwet tissue or pCi/kg wet tissue)

Dingsi = applied daily dose to species i from soil ingestion (pg/kg wet tissue/day orpCi/kg wet tissue/day)

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•i,soil = ingestion absorption factor for species i from soil ingestion (unitless)Ki = loss rate for species i, including depuration and metabolism (day-).

The applied daily dose to species i from soil ingestion is calculated from the equation(modified from EPA [1993]):

Divi = EC..,, SIj NIR,a, Fw , vW, (33)

whereDingsi = applied daily dose to species i from soil ingestion (pg/kg wet tissue/day or

pCi/kg wet tissue/day)ECmij = concentration in soil (ptg/kg dry or pCi/kg dry)SIi = soil ingestion rate of species i (kg soil ingested/kg dry diet)NIRtowi = total normalized ingestion rate for species i (kg prey wet tissue/kg

predator wet tissue/day). "Prey" may include plant species in thisdefinition.

Fdwi = conversion factor, dry diet to wet diet for species i (kg dry tissue/kg wettissue)

0i = area use factor for species i (ratio of contaminant area to home range -unitless)

Wi = seasonality factor for species i (fraction of year spent at the contaminatedsite - unitless)

The total normalized ingestion rate for species i is calculated from the equation (EPA1993):

NIRtai- w ( M E) (34)

j~i

whereNIRtotii = total normalized ingestion rate for species i (kg prey wet tissue/kg

predator wet tissue/day)FMRi = free-living metabolic rate of predator species i (kcal/day)Wi = body weight of species i (kg wet tissue)Pij -= wet weight or volume fraction of i's diet consisting -of preyj (unitless)MEj = metabolizable energy from preyj (kcal/kg prey wet tissue)

The metabolizable energy from preyj is calculated from (EPA 1993):

MEj = GE, AEJ (35)

where

MEj = metabolizable energy from preyj (kcal/kg prey wet tissue)GEj = gross energy from preyj (kcal/kg wet tissue)AEj = assimilation efficiency of preyj (unitless)

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Food Ingestion: The equilibrium contaminant tissue concentration in species i fromingestion of food is calculated from the following expression:

Cia = D (L (36)

whereCi,,g, = equilibrium tissue concentration in species i from ingestion of food (pg/kg

wet tissue or pCi/kg wet tissue)Dkgfi = applied daily dose to species i from food ingestion (pg/kg wet tissue/day

or pCi/kg wet tissue/day)Ctij = ingestion absorption factor for species i from ingestion of food speciesj

(unitless)Ki = loss rate of contaminant for species i, including depuration and

metabolism (1/day)

The applied daily dose to species i from food ingestion is calculated from the equation(modified from EPA [1993]):

Dbe =i V/ •.(Cj NIR,) (37)J

where

Dmng = applied daily dose to species i from food ingestion (pg/kg wet tissue/dayor pCi/kg wet tissue/day)

0i = area use factor for species i (ratio of contaminant area to home range -unitless)

Vi = seasonality factor for species i (fraction of year spent at the contaminatedsite - unitless)

Cj = average contaminant concentration inf' food item (pg/kg wet tissue orpCi/kg wet tissue)

NIRj = normalized ingestion rate oft" food type on a wet-weight basis (kg preywet tissue/kg predator wet tissue/day)

The normalized ingestion rate ofjf food type on a wet-weight basis is calculated (EPA1993) as:

NIRj = )j NIR,,,,o, (38)

whereNIRj = normalized ingestion rate ofjh food type on a wet-weight basis (kg prey

wet tissue/kg predator wet tissue/day)Pij wet weight or volume fraction of i's diet consisting of preyj (unitless)NIRtoti = total normalized ingestion rate for species i (kg prey wet tissue/kg

predator wet tissue/day)The term Pi is a large matrix describing which animal eats which plant or other animal.

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Total Terrestrial Animal Burden: The equilibrium tissue concentration for acontaminant in terrestrial animal species i is then calculated from the sum of the dermal,inhalation, and ingestion exposures. As can be seen from these equations, this modelrequires a great deal of information, including the absorption and depuration rates and thepredation (food source) matrix. Much of this can be approximated, or estimated usingallometric analyses (see below).

3.4.6 Allometric Models for Uptake by Animals

A key limitation to the use of transfer factors to relate intake of contaminants by variousanimal types to concentrations in animal products is the lack of measurements of variousradionuclides in most animals. A developing technique for estimating the transfer ofcontaminants in animal products is allometry (Higley, Domotor, and Antonio 2003).

The application of transfer factors for many types of animal products is problematicbecause of the wide range of organisms potentially exposed. For example, there are verylimited data available for riparian and terrestrial animals (i.e. very few terrestrialanimal:water and riparian animal:sediment concentration ratios). An alternative approach,called the kinetic/allometric method, was developed. Among other objectives, this workfilled in data gaps in the literature on transfer factors for specific riparian and terrestrialanimal receptors.

Allometry, or the biology of scaling, is the study of size and its consequences. There areseveral allometric equations that relate body size to many parameters, including ingestionrate, life span, inhalation rate, home range and more (Schmidt-Nielsen 1977; West et al.1997; Wilkie 1977). The most common form of allometric equation is the powerfunction

Y =a X (39)

where Y and X are size related measures, and a and f8 are constants. Allometric equationsprovide a way to estimate a biological process given a specific set of parameters. Whilethese equations were originally derived from empirical observations, there is a growingbody of evidence to suggest that these relationships have their origins in the dynamics ofenergy transport mechanisms. For example, allometry has been used to estimate the dailyrate of consumption by an animal in relation to its body size. Metabolic rate is known toscale to body mass to the / power (Calder 1984; Reiss 1989; West et al. 1997). Thisscaling has been found to extend to poikilothermic and ectothermic animals ( Bennett andDawson 1976) as well as unicellular ones (Wilkie 1977). Intraspecific scaling ofmetabolism has been found to vary somewhat, with the exponent of body weight largerfor juveniles than adults (Reiss 1989). However, allometric equations are meant torepresent qualitative trends over orders of magnitude of body mass. Consequently, thesegeneral allometric relationships can be used as general predictors regardless of age, orphyla to provide species-independent estimates of uptake.

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Higley, Domotor, and Antonio (2003) have developed equations to predict concentrationsof radionuclides in tissues for terrestrial and riparian animals exposed to contaminatedsoil, sediment or water. The equations were developed as a means to provide estimates oftransfer factors in the absence of data for particular species for radionuclides in soil,water or sediment. The allometric approach provides a potentially useful method toextrapolate radionuclide behavior across multiple species. While the method is simplistic,variations of this technique can be used to roughly estimate tissue concentrations,biological elimination rates, and uptake factors across a wide range of organisms. In theabsence of measured data it can provide a starting point for modeling environmentalexposure to contaminants.

3.5 Modeling the Effects of Food Processing on RadionuclideConcentrations

Changes in concentration in foods as a result of food preparation practices and activitiesare not commonly included in assessment models. Although such processes are routinelyacknowledged, there is a relative scarcity in the literature of information as to how theyshould be handled. Evidence exists that soaking, peeling, boiling, or baking caneffectively halve concentrations in consumed food. (NRPB 1995; Watterson andNicholson 1995; Jackson and Rickard 1998). For instance, the EPA Exposure FactorsHandbook does not even include a section about this process (EPA 1997). "Dosecalculations do not normally make allowance for such reductions, partly on the basis thatconsiderable variability may exist between individuals and that, for instance, water or fatsused for cooking may themselves be contaminated, so that any dose saving overall wouldbe marginal" (Jackson et al. 1998). In addition to that, there is a growing segment of thegeneral population that tend to reduce its consumption of the heavily processedfoodstuffs, as well as meat, poultry and other products.

Most radiological assessment codes use a simple approach to determining contaminantconcentrations in food crops; the concentrations in the crops are determined using asurface-deposition-to-concentration translocation factor for above-ground deposition, anda soil-to-edible-plant concentration ratio for growth in contaminated soils. This differssomewhat from the EPA approach to the same problems for dioxins and other lipophiliccontaminants (EPA 1994, EPA 1998). The EPA approach uses a soil-water-to-bulkcontamination ratio rather than a soil-to-bulk ration, and uses a Kd to relate soilcontamination to soil-water concentrations. A further assumption in the EPA method forthese contaminants is that these compounds do not transfer from contaminated surfaces(leaves or roots) into the vegetation, but are sorbed onto the surface. They then define an"empirical correction factor" (VG) that is essentially the ratio of the mass of the skin ofthe vegetable to the mass of the whole vegetable, further simplified with the assumptionthat the densities of the skin and bulk volume are the same, so that the ratio is of thevolume of skin to total volume of vegetable. This correction factor is then adjusted forwashing and peeling. The justification given in EPA (1994) is "Additional reductions inconcentration result from peeling, cooking, or cleaning, for example. Wipf, etal. (1982)found that 67% of unwashed carrot residues of 2,3,7,8-TCDD came out in wash water,and 29% was in the peels. A peeled, washed carrot correction factor might ... be, ... 0.04,

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(... 0.04 = 100% - 67% - 29%). A 96% reduction in the estimated VG for the potato (thepotato is cleaned and the skin is not eaten; additional reductions possibly when cookingthe potato) would equal 0.01. In a site-specific application, the type of vegetation,preparation, and so on, should be considered." The adjustment given by EPA is thus acombination of the washing and peeling and the assumption that nearly all contaminationis in the skin. This sort of correction factor should not be used with the soil-to-edible-plant concentration ratio models, because the CR used should already contain theassumption of low translocation and soil-to-plant transfer. Only the fraction of thereduction due to washing is always pertinent, and the peeling portion should beconsidered.

The NATO Manual FM-8-10-7 (Army 1993), Appendix F, gives decontaminationprocedures for foods contaminated by nuclear, chemical, and biological (NBC) agents.The procedures are designed for contamination resulting from NBC combat, and thedecontaminated foods are intended for consumption by soldiers or civilians in a post-combat situation; however, the processes and techniques are common ones that alsocould apply to home uses. As described in this manual, "some products can bedecontaminated by washing, peeling, or trimming the outer skin or leaves. Decontaminatepotatoes and hard-skinned fruits and vegetables by washing or scrubbing under runningwater, followed by peeling or scraping, then washing again. Potatoes, carrots, beets, andturnips can be washed at the supply depot. However, do not wash beans, rice, and onionsuntil they are delivered to the field kitchen; washing reduces their storage quality andshelf life. Citrus fruits, pineapples, corn, peas, beans, melons, pumpkins, cabbage, andnuts can be peeled. Decontaminate cucumbers, tomatoes, cherries, cranberries, grapes,pears, plums, and thin-skinned squash by soaking in a water or detergent solution andrinsing with vigorous agitation or brushing. Apricots, peaches, most berries, asparagus,broccoli, and leafy vegetables cannot be satisfactorily decontaminated because of fuzzysurfaces, irregular shapes, or small size which makes washing difficult. It is not practicalto decontaminate food contaminated through the food-chain. Meat and milk are the twomost common foodstuffs contaminated in this way.

"Milk may be decontaminated to a safe level by a complicated ion exchange process.The 1-131 activity will decline rapidly during storage of milk and milk-products, althoughthe Cesium and Strontium activity will remain almost constant for years. In an area withhigh-level fallout, milk is withdrawn from human consumption. The duration ofwithdrawal will be dependent upon the type of fallout and levels. Meat may bedecontaminated to a safe level by soaking in water or brine. Cesium is loosely bound inthe meat. By repeated soaking of meat cut in small pieces, most of the Cesium activitywill be removed. Traditional meat preserving, such as salting with brine, will remove upto 60 to 70 percent of the Cesium activity." (Army 1993). •

A fairly extensive review of available literature was conducted in association with theVAMP program of the IAEA (Noordijk and Quinalt 1992). This reference introduces afood processing retention factor, Fr, which is defined as the total amount of a contaminantin processed food divided by the total amount of contaminant in the original raw food.This report also introduces a related parameter, the processing efficiency P., which is the

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mass of prepared product divided by the mass of raw material. Numerous tables of thesevalues for various foods processed via various methods are presented. Averages andranges are presented in these tables. The ranges tend to be broad, and the average Frvalues appear to be around 0.5.

Noordijk and Quinalt (1992) stress that all Fr values referring to extraction proceduressuch as cooking, frying, etc., are only valid when the extraction liquid is removed and notused for other culinary purposes. The same applies to fractionation processes which takeplace, such as in the dairy industry or during the milling of grains. For example, the by-product of cheese making, whey, was formerly regarded as a waste or used to feedanimals. A large part of produced whey today is used as an additive for food for humans.Similarly, all fractions of the milling process are usually used to produce food. In thesecases, an Fr value of 1.0 should be used.

Noordijk and Quinalt (1992) summarize that the effects of processing on the behavior ofcontaminants depends on the contaminant, on the type of product, and on the method ofprocessing. Processing effects are rather small for root crops, whereas milling cerealgrains to flour will often remove about 70% of the contamination. The effects ofprocessing on vegetables and fruits are rather unpredictable, especially when thecontaminants are adsorbed on the surface of the plants.

Green and Wilkins (1995) extend the efforts of the IAEA study by Noordijk and Quinalt(1992). This review of the literature uses the same concepts (while renaming Fr as Rt)and adds additional references. This review also has reasonably detailed discussions ofthe applicability of the reported values for application to critical group analyses. Ingeneral, Green and Wilkins recommend values of 1.0 because of uncertainties in whetherliquids used in food processing would be consumed.

A possible method of handling the complexities of the food processing retention factorwas developed for this report. The method incorporates the uncertainties of both theretention in food of the contaminants, the variabilities in retention factor F, for differenttypes of cooking technique, as well as the uncertainties in whether the cooking liquors areconsumed. The concept can be described as:

Ceat, = C.a [F, + (I- F,.) S] (40)

where C.atn = the concentration in cooked or processed foods,Cra,, = the concentration in raw or unprocessed foods, andS = the fraction of cooking liquids and sauces that is consumed.

This formulation assumes that the contaminants lost in the food preparation process arereleased into the cooking liquids.

A complication to the process is that there are different types of food preparationprocesses. In evaluation of the data in the various reports, it appears that "dry cooking"techniques such as broiling, baking, or frying have different removal efficiencies than

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"wet cooking" techniques such as boiling, steaming, or stewing. In addition, differenttypes of foods will have different proportions of each preparation technique. These canbe accommodated by expanding the general equation for specific applications, as:

Ceate, = Cr. [ fdy Fray + (1 - Fa,.y ) Sdry} +

fwet ( F _we: + ( 1 - F,,,Wt) Sw, ) + fi.] (41)

In this formulation, the total intake of each type of food is subdivided into raw orunprocessed, wet cooked, and dry cooked. The approach may also be expanded toinclude other types of processing, as for milk, which includes fresh milk, cheeses, andwhey.

It seems apparent that this approach is impractical, because it has introduced 3 fractions,two processing retention factors, and a fractional consumption of cooking liquors foreach food type. It is unlikely that data will be available to populate such a model,although there might be data available in'the 1977 National Food Consumption Survey tobegin defining such parameters. However, in a stochastic model, wide ranges ofuncertainty may be combined into a single compound distribution.

A Crystal Ball® enabled EXCEL spreadsheet was used to evaluate the possibledistributions that might be encountered. The fractions of foods eaten in raw, wet-cooked,and dry-cooked states were broadly estimated by the author. The ranges for each of thepossibilities were assigned uniform distributions. Most are wide enough to accommodatenearly all eventualities; none of the results are overly sensitive to the assumptions. Thetotal fraction of foods consumed for each type was normalized to 1.0; this ensuresconservation of mass, but it also tends to minimize the sensitivity of the answer to thespecific distributional assumptions of the consumption fractions. The ranges of foodretention parameters are taken from Green and Wilkins (1995). Using the above generalequation, food processing retention factor distributions were generated for the commonfood types used in risk assessments: leafy vegetables, other vegetables, fruits, grains,meat, and milk.

The results of the computations confirm that for deterministic, maximum individual orcritical group calculations, a processing retention of 1.0 is a reasonable and not overly-conservative assumption. A value of 1.0 is always possible, and mean values for mostfood types are around 0.8 to 0.9. For leafy vegetables and other vegetables, the mean isabout 0.8, ranging from 0.4 to 1.0. For fruits and grains, the mean is nearly 0.95, and therange is well under a factor of 2. The range for meat is from 0.3 to 1.0, with a mean ofabout 0.8. Only for milk is there a substantial variability and non-linearity: the resultdepends heavily on how much cheese is eaten in proportion to milk, because the makingof cheese tends to reduce the contaminant concentrations.

As a result of the investigation, the distributions developed for the analysis are availablefor use. All of them are quite stable, except for the milk-processing factor. The resultsare summarized in Table 8. All distributions are simple - either uniform or triangular.Several approach right triangles, where the mode equals the maximum at 1.0. For milk, a

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triangular distribution is suggested, but the mode (most likely) value is essentiallyequivalent to the adjusted fraction of milk intake assigned to fresh milk; i.e., if 80 percentof milk intake is fresh, the mode of the triangle should be at 0.8, and if 50 percent of milkintake is fresh, the mode of the triangle should be at 0.5. This is the only parameter withsufficient sensitivity to warrant case-specific consideration in developing of exposurescenarios.

Table 8. Recommended Values of Food Processing Retention Factors for MostContaminants

Food Type Distribution Type Minimum Mode MaximumLeafy Uniform 0.5 0.85 1.0VegetablesOther Triangular 0.6 0.95 1.0VegetablesFruits Triangular 0.9 1.0 1.0Grains Uniform 0.9 - 1.0Meat Triangular 0.5 0.95 1.0Milk Triangular 0.25 0.8* 1.0* Distribution for milk is dependent on the fraction of milk intake assigned to cheese.The case here assumes 20 percent of milk intake results from consumption of locally-made cheese.

The radionuclide tritium (3H) is a special case. The food processing retention factors fortritium could be lower, because cooking tends to drive off water in foods, wherein mostof the tritium resides. The spreadsheet calculations described above for othercontaminants were modified and performed for tritium. The results are presented inTable 9. Note that there is relatively little difference from Table 1, indicating that it maynot be necessary to have any contaminant-specific inputs.

Table 9. Recommended Values of Food Processing Retention Factors for Tritium

Food Type Distribution Type Minimum Mode MaximumLeafy Uniform 0.5 0.85 1.0VegetablesOther Triangular 0.6 0.9 1.0VegetablesFruits Triangular 0.75 1.0 1.0Grains Uniform 0.99 - 1.0Meat Triangular 0.6 0.9 1.0Milk Triangular 0.5 1.0 1.0

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3.6 Modeling Other Terrestrial Exposure PathwaysIn the preparation of this report, a few unique examples of terrestrial human exposurepathways were identified. Although not part of standard food-chain pathways, theseadditional pathways deserve attention.

3.6.1 Unique Direct PathwaysThe pathway of showering with contaminated water, leading to inhalation exposure, iscommon to non-radiological assessments dealing with volatile chemical compounds (USEPA 1997); however, with the notable exception of radon, it is infrequently examined inradiological assessments because most radionuclides are not considered to be volatile.Even non-volatile contaminants can become entrained in water droplets and inhaled. Forsome radionuclides with very low GI-tract uptake, lung doses from this mechanism canbe higher than other organ doses from direct drinking of the water. The concentration ofradionuclides in the air of the shower can be related to the quantity of mist dropletsgenerated, using the absolute humidity of the air (Andelman 1990). As a general rule,radionuclide intakes via this pathway tend to be small. However, tritiated water cantranspire through the skin, and some models (e.g. GENII) sometimes multiply thenominal inhalation dose conversion factor by a factor of 1.5 to account for it (Napier etal. 1988).

A similar exposure pathway is Sweat bathing as part of the Sweat Lodge Ceremony, atraditional Native American custom (DOE 1997; Harris and Harper 1997). Based ontribal descriptions, between 0.5 and 3 hours/day on a weekly basis is assumed to be spentinside a sweat lodge kept at 60 - 80 degrees Centigrade (145 - 180 degrees Fahrenheit).No or very limited clothing is worn by the Ceremony participants. A large amount ofwood is used to heat up the rocks in an open outside fire adjacent to the Lodge. The sweatlodge is heated with these hot rocks, onto which water is poured to create steam. Airinside the sweat lodge is assumed to be saturated with water (equivalent to 0.3 kilogramsof water per m3 of air, and 0.3 L/m3 of semivolatiles and 2.5 L/m3 of volatiles), which arethen available for inhalation and dermal absorption over the entire body. During the 1hour of use, 4L of water is used. Doses resulting from intakes of some radionuclides viathis pathway can be larger than the associated doses via drinking water if theradionuclides have low GI-tract absorption, since this is a mechanism for inhalation oflarge amounts of water.

Evaporative coolers are commonly used in the hot, arid American West. Evaporativecoolers cool air by evaporating a stream of water into a stream of air. These coolersmight transfer water-borne contaminants to the indoor air. Thus, modeling must includean estimate of the radionuclide concentrations in indoor air when evaporative coolers arein operation so that the radiation dose for the human receptor who inhales thecontaminated air can be evaluated. Based on how evaporative coolers operate and theconservation of radioactivity (i.e., activity transferred to air is equal to the loss of activityfrom water), radionuclide concentrations in indoor air are estimated for this pathway byWu (2003) as

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Ma• = , Cw (42)

where

Cae, = activity concentration of radionuclide i in the air resulting from the operationof an evaporative cooler (Bq/m3)

f,. = fraction of radionuclides in water transferred to indoor air (dimensionless)Mwtjr = water evaporation rate (water use) for an evaporative cooler (m3/hr)Fai air flow rate for an evaporative cooler (m3/hr)Cw1 = activity concentration of radionuclide i in the groundwater (Bq/m3).

The fraction of radionuclides that remain in the reservoir, bleed-off water, or in the padsof evaporative cooler are not further modeled because it becomes a regular water sourcefor irrigation or sewage, or is removed when the pads are replaced, and therefore may notcause direct exposure (although for the case of continuous deposition of gamma emitterson the pad when the pad is replaced only once in every several years, there might besome concern of the long-term continuous direct exposure to the radionuclides depositedon the pad, depending on the location of the evaporative cooler inside the dwelling).Evaporation and air flow rates are estimated based on specifications of residentialevaporative cooling units. The typical evaporation rate is about 20 L/hr, and typical airflow rates range from 2,000 to 10,000 m3/hr (BSC 2003). The fraction of radionuclidestransferred from the water to the indoor air is an important parameter that is not availablein the literature, but the theoretical range is from 0 to 1 (BSC 2003). The calculation ofactivity concentrations in the air resulting from evaporative coolers does not includeconsideration of radionuclide buildup in the indoor air, because the air flow associatedwith the use of these coolers is such that during operation the windows of the buildingremain open and the number of air exchanges per hour is quite large. The calculatedindoor air concentrations are usually low because the air flow rate of the coolers is sohigh.

3.6.2 Unique Indirect PathwaysContamination of trees In Europe with 137Cs following the Chemobyl accident has led todetectable concentrations of radionuclides in wood. In villages in Ukraine and Belarus,wood is a common kitchen fuel. Wood ashes have been found to routinely exceed 50,000Bq/kg. Besides being a source of direct external exposure, the wood ashes were beingdistributed in kitchen gardens as soil amendments. The total administration of ashes inone year was equivalent to about 25% of the total 137Cs deposition on the garden from theinitial accident. Over a period of many years, this process resulted in reconcentration ofcesium and creation of an ingestion pathway (Dubreuil et al. 1999) that could bemultiples of 2 or 3 greater than the initial fallout soil contamination in the gardens.

Groundwater beneath the Hanford Site in eastern Washington State is contaminated withtritium at levels several times that of the drinking water standards (Poston, Hanf, andDirkes 2005). Trees near the Columbia River have roots that penetrate to thegroundwater. The trees have been found to not only contain tritium, but to transpire it tothe atmosphere. A similar source was documented at the Oak Ridge Site under Solid

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Waste Storage Area Five. In essence, the trees are a source of atmospheric release ofgroundwater tritium. However, the resulting atmospheric concentrations are low becauseof the large dispersion.

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4.0 Discussion

A number of alternative conceptual models for the key processes and mechanisms relatedto modeling the transport of radionuclides through the terrestrial biosphere have beenpresented. Several of the processes are not well known (e.g., foliar absorption andtranslocation, non-linear uptake of nutrient analogues, numerous transfer factors forunderstudied radionuclide plant/animal combinations) and offer areas for fruitfulresearch.

As a summary of the presentations, the following currently available techniques are themost likely to be successful in generic applications.

For soil, an annual average model incorporating:o Accumulation from irrigation or atmospheric deposition;o Leaching to deeper soil, using a Kd -modulated leach rate;o Harvest removal, averaged over crop typeso Uniform mixing in a reasonable (15-25 cm) soil surface layer, implicitly

caused by plowing, bioturbation, and leaching;o Neglecting radionuclide fixation, because appropriate long-term

measurements of distribution coefficient and concentration ratio willalready incorporate the effects;

o Neglecting surface-soil erosion losses, because eroded material from onelocation may accumulate in another, cancelling any perceived benefit.

" For resuspension, a mass-loading approach, because it has the lowest variabilityand is the most easily defended.

" Forfoliar interception, models incorporating:o Dry interception following the basic model suggested by Chamberlain

(1967) as updated by Adriano et al. 1982;o Wet interception considering one of the algorithms derived from the data

of Hoffman et al. (1989) or Prohl and Hoffman (1993)." For plant contamination with soil, an adhesion model applied so that material

translocated is not subject to weathering (e.g., IAEA 2003; Wu 2003)." For weathering, a single exponential model with a half-time between about 10

and 20 days, applied to material on the surface of the plant only." For translocation, a radionuclide-specific (or chemical-class-specific)

implementation allowing translocated materials to avoid being removed viaweathering processes (e.g., IAEA 2003; Wu 2003).

" For soil-to-plant transfer, a radionuclide- and plant-type-specific concentrationratio, where available, with:

o Site-specific application of additional crops, such as mushrooms andberries;

o The specific-activity model of Peterson and Davis (2002) for elementaland oxide forms of tritium;

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o A general specific-activity model for ' 4C in air, adapted to the RESRAD(Yu et al. 2001) model for transfer from irrigation water to air as 14C02;

o Perhaps site-specific application of specific-activity models for iodine andother micro-nutrient and macro-nutrient elements and their chemicalanalogues. This is a simplification of the need to allow for homeostaticregulation of some elements.

" Until a simple, robust Fruit Tree model is developed, continue to use theconcentration ratio approach for fruits.

* For animalproducts, a transfer factor approach applied to average daily feed,water, and soil intake by the animal where data are available.

o For other element/animal combinations, a justified mix of specific-activityand allometric models should be considered.

* For food processing, because the data are sparse and the reduction is generallysmall, modelers are justified in ignoring losses (which is equivalent to using afood-processing transmission factor of 1.0).

Site-specific models for additional little used or unusual pathways (e.g., honey,evaporative cooling systems) may need to be developed, justified, and used as appropriate forthe system and environment being analyzed.

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NRC FORM 335 U.S. NUCLEAR REGULATORY COMMISSION 1. REPORT NUMBER(9-2004) (Assigned by NRC, Add Vol.. Supp.. Rev.,NRCMD 3.7 and Addendum Numbers, If any.)

BIBLIOGRAPHIC DATA SHEET NUREG/CR-6910(See ist•.etons on ft re ,e,) PNNL-15872

2. TITLE AND SUBTITLE 3. DATE REPORT PUBUSHED

Alternative Conceptual Models for Assessing Foqd-Chain Pathways in Biosphere Models MONTH YEAR

June 2006

4. FIN OR GRANT NUMBER

Y6469

5. AUTHOR(S) 6. TYPE OF REPORT

B.A. Napier Technical

7. PERIOD COVERED (inclusive Dates)

8. PERFORMING ORGANIZATION - NAME AND ADDRESS (If NRC, provide Division, Ofic or Region, U.S. Nuclear Regulatory Commission. and mailing addrast; If contractor,provide name and mailing address.)

Pacific Northwest National LaboratoryRichland, WA 99352

9. SPONSORING ORGANIZATION - NAME AND ADDRESS (If NRC, Mme Same as above', IF contractor. provide NRC Division. Office or Region, U.S. Nuclear Regudatory Commission,and mealing address.)

Division of Fuel, Engineering, and Radiological ResearchOffice of Nuclear Regulatory ResearchU. S. Nuclear Regulatory CommissionWashington, DC 20555-0001

10. SUPPLEMENTARY NOTES

P.R. Reed, NRC Project Manager

11. ABSTRACT (200 words or loss)

This report describes alternative approaches to modeling the key processes in radionuclide transport in the biosphere. Becausethe focus of the NRC project Assessment of Food Chain Pathway Parameters in Biosphere Models is the food-chain modelsused in performance assessments of radioactive waste disposal facilities, models and approaches applicable over relatively longperiods (more than one year) are evaluated. There are a number of Important features and processes that all terrestrialbiosphere models must address: these include radionuclide behavior in soils, interception of deposition onto vegetation,'weathering of Intercepted material from plant surfaces, foliar absorption and translocation-within plants to other vegetativestructures, uptake from soil by plant roots, and transfer from plants to animals and animal products. Following a detaileddiscussion of possible alternatives, the author recommends approaches that should optimize the information gained for theresources expended. This food-chain pathway data may be used by the NRC staff to assess dose to persons in the referencebiosphere (e.g., persons who live and work in an area potentially affected by radionuclide releases) of waste disposal facilitiesand decommissioning sites.

12. KEY WORDSIDESCRIPTORS (List words orphrases that will assist researchers in iocaftg the report) 13. AVAILABILITY STATEMENT

terrestrial biosphere modeling unlimitedalternative conceptual models 14. SECURITY CLASSIFICATION

transfer factors Mis Page)

uptake factors unclassifiedinterception fraction (This Rep•r

weathering and retention unclassified15. NUMBER OF PAGES

16. PRICE

NR FRM3l(920..PINEDO REYLD AENRC FORM 335 (9-2004) PRINTED ON RECYCLED PAPER

Federal Recycling Program